United States

Environmental Protection Agency

Office of Water

4304

EPA-822-B-01-003

October 2001

Nutrient Criteria

Technical Guidance Manual

Estuarine and Coastal

Marine Waters
U.S. Environmental Protection Agency

Nutrient Criteria

Technical Guidance Manual

Estuarine and Coastal Marine Waters

October 2001
ii Nutrient Criteria—Estuarine and Coastal Waters

Disclaimer

This manual provides technical guidance to States, Indian Tribes, and other authorized jurisdictions to

establish water quality criteria and standards under the Clean Water Act (CWA), to protect aquatic life

from acute and chronic effects of nutrient overenrichment. Under the CWA, States and Indian Tribes are

to establish water quality criteria to protect designated uses. State and Indian Tribal decisionmakers

retain the discretion to adopt approaches on a case-by-case basis that differ from this guidance when

appropriate and scientifically defensible. Although this manual constitutes EPA's scientific

recommendations regarding ambient concentrations of nutrients that protect resource quality and aquatic

life, it does not substitute for the CWA or EPA's regulations; nor is it a regulation itself. Thus, it cannot

impose legally binding requirements on EPA, States, Indian Tribes, or the regulated community, and

might not apply to a particular situation or circumstance. EPA may change this guidance in the future.

Cover Photograph: Somewhere on the Chesapeake Bay. Supplied by David Flemer as a duplicate copy

from the Chesapeake Biological Laboratory Photo Archives, University of Maryland; date unknown but

earlier than 1972. 
Nutrient Criteria—Estuarine and Coastal Waters iii

CONTENTS

Contributors ........................................................................ xi

Acknowledgments .................................................................. xiii

Foreword ......................................................................... xv

Executive Summary ................................................................ xvii

Chapter 1. Introduction and Objectives ............................................... 1-1

1.1 Backround ..................................................................... 1-1

1.2 Definition of Estuaries and Coastal Systems .......................................... 1-2

1.3 Nature of the Nutrient Overenrichment Problem in Estuarine and Coastal Marine Waters ...... 1-5

Scope and Magnitude of the Problem ............................................... 1-5

1.4 The Nutrient Criteria Development Process ........................................... 1-8

Preliminary Steps .............................................................. 1-8

Strategy for Reducing Human-Based Eutrophication ................................. 1-10

Nutrient Criteria Development Process ............................................ 1-11

Chapter 2. Scientific Basis for Estuarine and Coastal Waters Quantitative Nutrient Criteria . . 2-1

2.1 Introduction .................................................................... 2-1

Purpose and Overview .......................................................... 2-1

Some Important Nutrient-Related Scientific Issues .................................... 2-2

River-to-Ocean Continuum: Watershed/Nearshore Coastal Management Framework ......... 2-5

2.2 Controlling the Right Nutrients ................................................... 2-10

Overview .................................................................... 2-10

Some Empirical Evidence for N Limitation of Net Primary Production ................... 2-11

Some Threshold Responses to Nitrogen Overenrichment .............................. 2-13

Effects of Physical Forcing on Net Primary Production ............................... 2-13

Other Physical Factors ......................................................... 2-23

2.3 Nutrient Loads and Concentrations: Interpretation of Effects ............................ 2-21

Conceptual Framework ......................................................... 2-24

Examples .................................................................... 2-24

2.4 Physical-Chemical Processes and Dissolved Oxygen Deficiency ......................... 2-27

2.5 Nutrient Overenrichment Effects and Important Biological Resources .................... 2-28

Benthic Vascular Plant Responses to Nutrients ...................................... 2-28

Other Examples of Important Biotic Effects of Nutrient Overenrichment .................. 2-29

2.6 Concluding Statement on Nitrogen and Phosphorus Controls ............................ 2-32

Chapter 3. Classification of Estuarine and Coastal Waters ............................... 3-1

3.1 Introduction .................................................................... 3-1

Purpose and Background ........................................................ 3-1

Defining the Resource of Concern ................................................. 3-2

3.2 Major Factors Influencing Estuarine Susceptibility to Nutrient Overenrichment .............. 3-2

Dilution ...................................................................... 3-3

Water Residence Time .......................................................... 3-3

Stratification .................................................................. 3-3
CONTENTS (continued)

iv Nutrient Criteria—Estuarine and Coastal Waters

3.3 Examples of Coastal Classification ................................................. 3-4

Geomorphic Classification ....................................................... 3-4

Man-Made Estuaries ............................................................ 3-6

Physical/Hydrodynamic Factor–Based Classifications ................................. 3-6

Other Considerations ........................................................... 3-9

Summary .................................................................... 3-11

3.4 Coastal Waters Seaward of Estuaries ............................................... 3-11

Geomorphic Classification ...................................................... 3-12

Nongeomorphic Classification ................................................... 3-13

Chapter 4. Variables and Measurement Methods To Assess and Monitor

Estuarine/Marine Eutrophic Conditions .......................................... 4-1

4.1 Introduction .................................................................... 4-1

4.2 Causal and Response Indicator Variables ............................................. 4-2

Nutrients as Causal Variables ..................................................... 4-2

Response Variables ............................................................. 4-6

Measures of Water Clarity ....................................................... 4-7

Dissolved Oxygen .............................................................. 4-8

Benthic Macroinfauna .......................................................... 4-9

4.3 Field Sampling and Laboratory Analytical Methods ..................................... 4-9

Field Sampling Methods ......................................................... 4-9

Laboratory Analytical Methods .................................................. 4-13

Water Column Nutrients ........................................................ 4-13

Sediment Analyses ............................................................ 4-15

Determination of Primary Productivity ............................................ 4-16

Phytoplankton Species Composition .............................................. 4-16

Macrobenthos, Macroalgae, and Seagrasses and SAV ................................. 4-17

Chapter 5. Databases, Sampling Design, and Data Analysis .............................. 5-1

5.1 Introduction .................................................................... 5-1

5.2 Developing Regional and National Databases for Estuaries and Coastal Waters .............. 5-1

Data Sources .................................................................. 5-3

EPA Water Quality Data ........................................................ 5-3

National Oceanographic and Atmospheric Administration (NOAA) ...................... 5-5

Rivers and Streams Water Quality Data ............................................. 5-6

USGS San Francisco Bay Program ................................................. 5-6

State/Tribal Monitoring Programs ................................................. 5-7

Sanitation Districts ............................................................. 5-7

Academic and Literature Sources .................................................. 5-8

Volunteer Monitoring Programs ................................................... 5-8

Quality of Historical Data ........................................................ 5-9

Location Data ................................................................. 5-9

Variables and Analytical Methods ................................................. 5-9

Laboratory Quality Control ...................................................... 5-9

Data Collecting Agencies ....................................................... 5-10

Time Period ................................................................. 5-10
CONTENTS (continued)

Nutrient Criteria—Estuarine and Coastal Waters v

Index Period ................................................................. 5-10

Representativeness ............................................................ 5-10

Gathering New Data ........................................................... 5-10

5.3 Sampling Design ............................................................... 5-10

Sampling Protocol ............................................................. 5-11

Sampling Technique ........................................................... 5-12

Initial Considerations .......................................................... 5-12

Specifying the Population and Sample Unit ......................................... 5-13

Specifying the Reporting Unit ................................................... 5-14

Sources of Variability .......................................................... 5-14

Alternative Sampling Designs ................................................... 5-16

Monitoring Programs .......................................................... 5-18

Citizen Monitoring Programs .................................................... 5-20

5.4 Quality Assurance/Quality Control ................................................ 5-21

Representativeness ............................................................ 5-21

Completeness ................................................................ 5-21

Comparability ................................................................ 5-21

Accuracy .................................................................... 5-22

Variability ................................................................... 5-22

5.5 Statistical Analyses ............................................................. 5-22

Data Reduction ............................................................... 5-22

Frequency Distributions ........................................................ 5-23

Correlation and Regression Analyses ............................................. 5-23

Tests of Significance .......................................................... 5-24

Chapter 6. Determining the Reference Condition ....................................... 6-1

6.1 Introduction and Definition ....................................................... 6-1

6.2 Significance of Reference Conditions ............................................... 6-1

6.3 Paucity of Similar Estuarine and Coastal Marine Ecosystems ............................. 6-4

6.4 Approaches for Establishing Reference Conditions ..................................... 6-4

In Situ Observations as the Basis for Estuarine Reference Condition ...................... 6-5

Areal Load Approach to Identification of Reference Condition ......................... 6-11

Chapter 7. Nutrient and Algal Criteria Development .................................... 7-1

7.1 Introduction .................................................................... 7-1

7.2 Role of Regional Technical Assistance Groups ........................................ 7-2

7.3 Classification .................................................................. 7-3

7.4 Descriptive Background Information ................................................ 7-3

Estuarine Watershed Characterization .............................................. 7-3

Within Estuarine System Characterization ........................................... 7-4

7.5 Elements of Nutrient Criteria ...................................................... 7-5

Reference Condition ............................................................ 7-5

Historical Information .......................................................... 7-5

Models ...................................................................... 7-5

Antidegradation Policy and Attention to Downstream Effects ........................... 7-6

The RTAG ................................................................... 7-7
CONTENTS (continued)

vi Nutrient Criteria—Estuarine and Coastal Waters

7.6 Hypothetical Examples of Nutrient Criteria Development Deliberations .................... 7-7

Scenario ..................................................................... 7-7

7.7 Evaluation of Proposed Criteria .................................................... 7-8

Guidance for Interpreting and Applying Criteria ...................................... 7-9

Do the Criteria Protect Designated Uses? ........................................... 7-9

Restoration Goals ............................................................. 7-11

Sampling for Comparison to Criteria ........................................... 7-11

7.8 Nutrient Criteria Interpretation Procedures .......................................... 7-12

Decisionmaking Protocol ....................................................... 7-12

Multivariable Enrichment Index .................................................. 7-12

Frequency and Duration ........................................................ 7-13

7.9 Criteria Modifications ........................................................... 7-14

7.10 EPA, State, or Tribe Responsibility under the Clean Water Act ......................... 7-14

7.11 Implementation of Nutrient Criteria into Water Quality Standards ....................... 7-14

Chapter 8. Using Nutrient Criteria to Protect Water Quality ............................. 8-1

8.1 Managing Point Source Pollution ................................................... 8-1

The Clean Water Act and Water Quality Standards .................................... 8-1

Protecting Designated Uses ...................................................... 8-2

Maintaining Existing Water Quality ................................................ 8-3

General Policies ............................................................... 8-4

Providing Flexibility in Implementation ............................................. 8-6

NPDES Permits ................................................................ 8-7

Look to the Future ... Pollutant Trading ............................................ 8-10

8.2 Managing Nonpoint Source Pollution .............................................. 8-11

Nonpoint Sources of Nutrients ................................................... 8-12

Efforts to Control Nonpoint Source Pollution ....................................... 8-13

National Estuary Program ....................................................... 8-14

Atmospheric Deposition ........................................................ 8-15

Coastal Nonpoint Pollution Control Programs ....................................... 8-16

Farm Bill Conservation Provisions ................................................ 8-17

8.3 Comprehensive Procedure for Nutrient Management .................................. 8-19

Step 1: Status Identification ..................................................... 8-19

Step 2: Background Investigation ................................................ 8-20

Step 3: Data Gathering and Diagnostic Monitoring .................................. 8-21

Step 4: Source Identification .................................................... 8-23

Step 5: Management Practices for Nutrient Control .................................. 8-24

Step 6: Detailed Management Plan Development .................................... 8-26

Step 7: Implementation and Communication ....................................... 8-26

Step 8: Evaluation Monitoring and Periodic Review ................................. 8-26

Step 9: Completion and Evaluation ............................................... 8-27

Step 10: Continued Monitoring of the System ...................................... 8-27

8.4 Resources .................................................................... 8-28

Chapter 9. Use of Models in Nutrient Criteria Development .............................. 9-1

9.1 Introduction .................................................................... 9-1

Use of Empirical Models in Nutrient Criteria Development ............................. 9-2
CONTENTS (continued)

Nutrient Criteria—Estuarine and Coastal Waters vii

Use of Mathematical Models in Nutrient Criteria Development .......................... 9-3

9.2 Model Identification and Selection .................................................. 9-3

Model Identification ............................................................ 9-4

Model Selection ............................................................... 9-8

9.3 Model Classification ............................................................. 9-9

Level I Models ............................................................... 9-10

Level II Models ............................................................... 9-13

Level III Models .............................................................. 9-14

Level IV Models .............................................................. 9-15

Summary of Model Capabilities .................................................. 9-17

9.4 Use of Models for Nutrient Investigation ........................................... 9-17

Model Calibration and Validation ................................................ 9-17

9.5 Management Applications ....................................................... 9-22

Load-Response Analysis ........................................................ 9-22

Acceptable Nutrient Loads ...................................................... 9-23

Case Study Example ........................................................... 9-24

References ....................................................................... R-1

Appendixes

Appendix A: Conditions for Bloom Development: Interplay among Biogeochemical, Biological, 

and Physical Processes ................................................ A-1

Appendix B: Additional Information on the Role of Temperature and Light on Estuarine 

and Coastal Marine Phytoplankton ....................................... B-1

Appendix C: Additional Information on Flushing in Estuaries ............................ C-1

Appendix D: NOAA Scheme for Determining Estuarine Susceptibility ..................... D-1

Appendix E: Comparative Systems Empirical Modeling Approach: the Empirical Regression 

Method to Determine Nutrient Load-Ecological Response Relationships 

for Estuarine and Coastal Waters ........................................ E-1

Appendix F: Selected Theoretical Approaches to Classification of Estuaries and Coastal Waters . . F-1

Appendix G: Examples of Nutrient Concentration Ranges and Related Hydrographic Data 

for Selected Estuaries and Coastal Waters in the Contiguous States 

of the United States ................................................... G-1

Appendix H: Preliminary Statement of Proposed Near Coastal Marine Nutrient Sampling and

Reference Condition Development Procedure .............................. H-1

Case Studies

San Francisco Bay Program: Managing Coastal Resources of the U.S. ....................... CS-1

Long Island Sound - Hypoxia ....................................................... CS-4

NP Budget for Narragansett Bay .................................................... CS-15

Tampa Bay Case Study ........................................................... CS-19

Restoring Chesapeake Bay Water Quality ............................................. CS-27

A Perspective from Washington State ................................................ CS-42
CONTENTS (continued)

viii Nutrient Criteria—Estuarine and Coastal Waters

Figures

Figure 1-1a. Draft aggregation of Level III ecoregions for the National Nutrient Strategy illustrating

those areas most related to coastal and estuarine nutrient criteria development. ....... 1-3

Figure 1-1b. Coastal provinces ....................................................... 1-4

Figure 1-2. The eutrophication process. ............................................... 1-6

Figure 1-3. Expanded nutrient enrichment model Source: Bricker et al. 1999. ................. 1-7

Figure 1-4. Elements of nutrient criteria development and their relationships in the process. ...... 1-9

Figure 1-5. Derivation of the reference condition and the National Nutrient Criteria Program 

using TP, TN, and Chlorophyll a as example variables. .......................... 1-9

Figure 1-6. Flowchart of the nutrient criteria development process. ......................... 1-12

Figure 2-1. Idealized scheme defining the coastal ocean and the coastal zone ................. 2-2

Figure 2-2. Schematic representation of contemporary (Phase II) conceptual model of coastal

eutrophication. ......................................................... 2-5

Figure 2-3. Salinity zones .......................................................... 2-7

Figure 2-4. Schematic illustrating the central role of phytoplankton as agents of biogeochemical

change in shallow coastal ecosystems. ....................................... 2-7

Figure 2-5. Transport of nutrients to Laholm Bay, Sweden ............................... 2-12

Figure 2-6. Summary of nitrogen:phosphorus ratios in 28 sample estuarine ecosystems ......... 2-14

Figure 2-7. Factors that determine whether nitrogen or phosphorus is more limiting in 

aquatic ecosystems ..................................................... 2-15

Figure 2-8. Cartoon diagrams of three physical forcings that operate at the interface between SCEs 

and the coastal ocean (tides), watershed (river inflow), and atmosphere (wind) ...... 2-19

Figure 2-9. Simple schematic diagram showing the influences of river flow on ecosystem 

stocks and processes examined in this study. ................................. 2-20

Figure 2-10. Scatter diagram showing the relationship between the rate of decline in 

dissolved-oxygen concentrations in deep water and average deposition rates 

of total chlorophyll a during the spring-bloom period. .......................... 2-20

Figure 2-11 Schematic diagram of coastal plain estuary types, indicating direction and 

a-d. degree of mixing. ...................................................... 2-22

Figure 2-12. Net transports in estuaries resulting from estuarine flows and mixing. ............. 2-23

Figure 2-13. Net movement of a particle in each layer of a two-layered flow system. ............ 2-23

Figure 2-14. The fraction of landside nitrogen input exported from 11 North American and 

European estuaries versus freshwater residence time (linear time scale). ........... 2-25

Figure 2-15. Scatter plots of water column averaged chlorophyll a at a mesohaline station versus 

several different functions of total nitrogen loading rate measured at the fall line of 

the Potomac River estuary ............................................... 2-26

Figure 2-16a. Primary production by phytoplankton as a function of the estimated annual input of

dissolved inorganic nitrogen per unit volume of a wide range of marine ecosystems. . 2-30

Figure 2-16b. Primary production by phytoplankton as a function of the annual input of dissolved

inorganic nitrogen per unit area of a wide range of marine ecosystems. ............ 2-30

Figure 2-16c. Fisheries yield per unit area as a function of primary production in a wide range of

estuarine and marine systems. ............................................. 2-31

Figure 2-17. Comparative evaluation of fishery response to nutrients. ........................ 2-32

Figure 3-1. Idealized micronutrient-salinity relations showing concentration and mixing of 

nutrient-rich river water with nutrient-poor seawater ........................... 3-10

Figure 3-2a. Relationship between the mean annual loadings of dissolved inorganic nitrogen 

and the mean annual concentration of chlorophyll a in microtidal and macrotidal 

estuaries .............................................................. 3-11
CONTENTS (continued)

Nutrient Criteria—Estuarine and Coastal Waters ix

Figure 3-2b. Relationship between the mean annual concentrations of dissolved inorganic 

nitrogen and chlorophyll a in microtidal and macrotidal estuaries ................ 3-12

Figure 6-1. Environmental quality scale representing reference conditions and potential nutrient

criteria relative to designated uses .......................................... 6-2

Figure 6-2. Hypothetical frequency distribution of nutrient-related variables showing quantities for

reference or high-quality data and mixed data. ................................. 6-7

Figure 6-3. Hypothetical example of load/concentration response of estuarine biota to increased

enrichment ............................................................. 6-9

Figure 6-4. An illustration of the comparison of past and present nutrient data to establish 

a reference condition for intensively degraded estuaries. ........................ 6-10

Figure 6-5. Areal load estimate approach to nutrient reference condition determination. ........ 6-14

Figure 7-1. Generalized progression and relationship of the elements of a nutrient criterion ...... 7-2

Figure 7-2. Hypothetical illustration of developing a TN criterion in an estuary ................ 7-8

Figure 8-1. Components of water quality standards ...................................... 8-2

Figure 8-2. “Threefold framework” of evaluation. ...................................... 8-25

Figure 9-1. Eutrophication model framework ........................................... 9-6

Figure 9-2. Use of models in load-response analysis .................................... 9-23

Figure 9-3. Use of models in determining allowable loads ................................ 9-24

Figure 9-4. Shipps Creek site map and salinity monitoring locations ........................ 9-25

Figure 9-5. Model results for existing conditions ....................................... 9-27

Figure 9-6. Model results for 50% reduction in WWTP load .............................. 9-28

Tables

Table 2-1. Categorization of the world’s continental shelves based on location, major river, 

and primary productivity .................................................. 2-8

Table 2-2. Estuaries exhibiting seasonal shifts in nutrient limitation with spring P limitation 

and summer N limitation ................................................. 2-12

Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL

mesocosm enrichment experiment ......................................... 2-16

Table 3-1. General drowned river valley estuarine characteristics .......................... 3-5

Table 3-2. Classification of coastal systems based on relative importance of river flow, tides, 

and waves to mixing ..................................................... 3-9

Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of 

coastal and marine environments .......................................... 4-10

Table 6-1. Summary of estuarine and coastal nutrient reference condition determinations ....... 6-5

Table 6-2. Requisite assumptions for establishing watershed-based reference conditions ....... 6-12

Table 7-1. Example of an enrichment index using the middle portion of a hypothetical estuary . . 7-13

Table 8-1. States for which the nonpoint source agency is not the water quality agency ........ 8-14

Table 8-2. States and territories with coastal nonpoint pollution control programs ............ 8-17

Table 9-1. Basic model features .................................................... 9-18

Table 9-2. Key features of selected models ........................................... 9-19

Table 9-3. Calculation spreadsheet for Shipps Creek estuary ............................. 9-26

Table 9-4. Chesapeake watershed nitrogen deposition under varying management schemes for

emissions of nitrogen atmospheric depositions precursors ....................... 9-29

Table 9-5. Water quality state variables used in CBEMP ................................ 9-31
x Nutrient Criteria—Estuarine and Coastal Waters
Nutrient Criteria—Estuarine and Coastal Waters xi

CONTRIBUTORS

Jessica Barrera (Hispanic Association of Colleges and Universities/University of Miami) 

Robert Cantilli (U.S. Environmental Protection Agency)

Ifeyinwa Davis (U.S. Environmental Protection Agency)*

Edward Dettmann (U.S. Environmental Protection Agency) 

Jen Fisher (University of Georgia)

David Flemer (U.S. Environmental Protection Agency)*

Thomas Gardner (U.S. Environmental Protection Agency)*

George Gibson (U.S. Environmental Protection Agency)*

Debbi Hart (U.S. Environmental Protection Agency)

James Latimer (U.S. Environmental Protection Agency)

Scott Libby (Battelle)*

Greg Smith (GLEC, Inc.)

CarolAnn Siciliano (U.S. Environmental Protection Agency)*

Jack Word (MEC Analytical Systems)*

*Denotes primary authors

WORK GROUP MEMBERS

Ifeyinwa Davis (U.S. Environmental Protection Agency)

David Flemer (U.S. Environmental Protection Agency)

John Fox (U.S. Environmental Protection Agency)

George Gibson (U.S. Environmental Protection Agency)

Debbi Hart (U.S. Environmental Protection Agency)

Suzanne Bricker (National Oceanic and Atmospheric Administration)

Dorothy Leonard (National Oceanic and Atmospheric Administration)

Scott Libby (Battelle)

Greg Smith (GLEC, Inc.)

Jack Word (MEC Analytical Systems)
xii Nutrient Criteria—Estuarine and Coastal Waters
Nutrient Criteria—Estuarine and Coastal Waters xiii

ACKNOWLEDGMENTS

The authors wish to thank the following peer reviewers for their assistance in the preparation of this

manual: Don Boesch (University of Maryland) and Hans Paerl (Institute of Marine Sciences, University

of North Carolina), Matthew Liebman (U.S. EPA Region I), Michael Bira (U.S. EPA Region VI), and

Kenneth Teague (U.S. EPA Region VI). Walter Nelson and Peter Eldridge (U.S. EPA Office of

Research and Development) provided comments on an early draft of the manuscript, and Thomas

Brosnan (National Oceanic and Atmospheric Administration, Damage Assessment Center), Michael

Kemp (Horn Point Laboratory, CES, University of Maryland), Jonathan Pennock (Dauphin Island

Laboratory, University of South Alabama), Hassan Mirsajadi (Delaware Department of Natural

Resources and Environmental Control), and David Tomasko (State of Florida Southwest Florida Water

Management District) provided formal peer review comments on a final working draft. Additional

comments on the working draft were provided by Suesan Saucerman (U.S. EPA Region IX); John (Jack)

Kelly, James Latimer, and Edward Dettmann (U.S. EPA ORD); Lewis Linker (U.S. EPA Chesapeake

Bay Program); Laura Gabanski (U.S. EPA Office of Wetlands, Oceans, and Watersheds); Joel Salter

(Office of Wastewater Management); Mimi Dannel and Marjorie Wellman (U.S. EPA Office Science

and Technology); CarolAnn Siciliano (U.S. EPA Office of General Council); and Cynthia Moncreiff

(University of Southern Mississippi). Treda Smith (U.S. EPA Office of Water) assisted in compiling

references. Edits and suggestions made by the peer review panel were incorporated into the final version

of the manual. State agencies and private interest groups also offered comments and they were addressed

where possible in this manual. We appreciate the work of Joanna Taylor, The CDM Group, Inc., who

patiently and graciously made repeated format changes to this manuscript as it evolved over the several

months. The authors of the case studies are thanked for their contributions.

Estuaries and coastal waters are a diverse suite of ecosystems, and differences in methods and

approaches are to be expected. This and subsequent manuals are not intended to be singular, one-time

publications. As experience accumulates, future editions will be prepared to reflect new understanding.
xiv Nutrient Criteria—Estuarine and Coastal Waters
Nutrient Criteria—Estuarine and Coastal Waters xv

FOREWORD

This manual is intended for State/Tribal and Federal agency personnel actively engaged in water resource

management data collection, assessment, planning, and project implementation. Consequently, it

incorporates both a scientific rationale and enough of the “nuts and bolts” of nutrient criteria

development and management to help both initiates and those experienced in water resource

management.

These nutrient criteria development and management efforts are directed at anthropogenic sources. 

Inherent “natural” background levels are not and should not be subject to management. Our

responsibility is to abate human-caused eutrophication in estuaries and coastal or “near coastal” (out to

about 20 nautical miles) marine waters. 

To distinguish between natural background enrichment and human impacts, it is necessary to identify

localities that experience minimal human influence. Ambient nutrient measurements at these sites may

then be compared to similar sites that do experience human influences. The difference in nutrient

measurements is the difference between a reference site and a test site. A reference condition is a

collection of measurements from several reference sites that incorporates a central tendency statistic.

Because of differences in geologic parent material, climate, and geography, reference conditions are

different from one region to another. Similarly, waterbodies, especially estuaries, often respond

differently to nutrient inputs. Lakes and reservoirs are different from streams and rivers, and estuaries

and coastal marine waters have characteristics different from both. Criteria have to be designed for

particular waterbody types and the regions in which they lie.

The primary variables of concern in criteria development are two causal enrichment variables: total

phosphorus (TP) and total nitrogen (TN). These nutrients are essential to algal and plant production and

are the base of the food chain that supports all other life in the system. Also, two initial response

variables usually are the first indicators of biological growth reaction to enrichment. One is chlorophyll

a, which indicates photosynthesis and biomass production; the other is Secchi depth, a measure of water

clarity or a measure of turbidity, reflecting planktonic growth in the absence of inorganic suspended

material. In many marine and estuarine instances dissolved oxygen concentration (DO) and macrophyte

growth and density are also important measures and, where indicated, may be included as initial response

variables. Other measures can also be used, but these have been selected by EPA as of primary concern.

Nutrient criteria consist of judicious incorporation of present reference condition information about the

primary variables, together with a knowledge of historical conditions and trends in the nutrient quality

of the resource. These two factors, possibly augmented by data extrapolations or models, are analyzed

objectively by a panel of regional specialists well versed in the biology, physics, and chemistry of the

systems of concern. The criteria are also evaluated with respect to the possible consequences of their

implementation on downstream receiving waters. All of these elements are required for the

development of a nutrient criterion.
xvi Nutrient Criteria—Estuarine and Coastal Waters

With this information, the status of a given water resource can be determined, management plans can be

made, and management efforts can be evaluated.

The best possible understanding of the physical, chemical, and biological interrelationships in the

environment is important in nutrient criteria development and the subsequent management response. 

However, effective nutrient criteria can and should be developed even in the absence of an in-depth

scientific investigation of the ecological processing of nutrients in the estuarine and marine environment. 

An adequate number of proximal reference sites and current knowledge of the system are sufficient to

initiate criteria development and proposed management responses. A conservative, environmentally

responsible start can be made toward alleviating nutrient pollution, subject to adjustment as more

scientific knowledge is obtained and verified. 

The reference condition approach to criteria development was peer reviewed by the USEPA Science

Advisory Board in 1990 and 1994 and judged to be scientifically defensible. It is also likely to be the

most cost-effective approach.
Nutrient Criteria—Estuarine and Coastal Waters xvii

EXECUTIVE SUMMARY

This manual is designed for use by State, Tribal, and Federal water resource managers as they address the

cultural enrichment of their waters in conjunction with the EPA National Nutrient Criteria Program. It is

intended to provide a stepwise sequence of actions leading to the development of nutrient criteria for

estuarine and near-coastal marine waters to be used in correcting this overenrichment problem.

The premise of the National Nutrient Criteria Program is that many, if not most, of our nation’s estuarine

and coastal waters are moderately to severely polluted by excessive nutrients (Bricker et al.1999),

especially nitrogen and phosphorus. This nutrient pollution affects not only the biotic integrity of the

waters and the decline of valuable fish and shellfish, it has the potential to cause harm to the public

health through hazardous algal blooms and the propagation of waterborne diseases. To address this

problem, EPA uses a regionalized, waterbody type specific approach to the development of nutrient

criteria or benchmarks for management decisionmaking. These criteria are based on the measurement of

the most natural (or least impacted by human development) waters of a given type in a given area

reflecting the condition to be expected in that region if human impacts are not a factor or are at least

minimized. The variables of specific concern are total phosphorus and total nitrogen as causal variables,

algal biomass (e.g., chlorophyll a for phytoplankton and ash-free dry weight for macroalgae), and water

clarity (e.g., Secchi depth) as early response variables. In waters that already experience hypoxia,

dissolved oxygen should be added as a response variable. EPA encourages States and Tribes to consider

additional response indicators such as seagrasses and algal species composition.

This natural ambient background or “reference condition” is an important element of the nutrient criteria

to be developed. The other elements are: an understanding of the historical status and trend of the water

resource to help put the reference condition in perspective; models of the nutrient data to help better

understand historical and present information and to project future consequences; concern and attention

to the effects of any criteria development on downstream receiving waters; and the objective compilation

and assessment of all of this information by a skilled body of regional experts...the “Regional Technical

Assistance Group” or RTAG. The regional criteria so developed are guidelines the States and Tribes of

the continental United States can use as they prepare their own criteria and standards for the

improvement and protection of the nation’s coastal waters.

The first of the actions needed to reach this criteria objective is the organization and utilization in each

EPA Region of an RTAG consisting of specialists from State and Federal natural resource management

agencies versed in the management and scientific principles most appropriate to that region and those

waters. These are water resource managers, oceanographers, chemists, land use specialists, biologists,

estuarine ecologists, statisticians, and similar local civil service experts employed by the State or Federal

government. Academicians, special interest groups, and environmental group representatives are also

important participants in the criteria development process and may assist the RTAG in its efforts.

The first requirement of the RTAG is the review and refinement of ecoregional determinations as 

most appropriate to the area. These are the geographic boundaries surrounding the similar estuarine and
xviii Nutrient Criteria—Estuarine and Coastal Waters

coastal marine waters for which the criteria will apply. They are based on the EPA Ecoregion concept

and incorporate attendant coastal Provinces, both of which are based on geographic and geologic

similarities of landforms and parent material. The importance of this regionalization is the effort to deal

with waters all having a similar inherent background nutrient loading and response characteristic. Once

the regional boundaries and perhaps subregional divisions are completed, the RTAG investigates the

physical classification of the waters into similar estuaries or coastal reaches or embayments for criteria

development. In many instances the estuaries may be unique and require specific criteria.

Within the classification scheme developed, reference sites are identified as those areas suffering the

least cultural development or impact, and the compilation of similar reference sites becomes a reference

condition. The manual describes the scientific rationale for the variables selected, the dynamics of the

receiving waters, and potentially confounding physical and chemical interrelationships influencing

criteria development. It also describes sampling and analytical techniques for data gathering and

processing to develop the reference conditions as well as several options for the compiling of this

information. These include: (1) recognition and measurement of an excellent water body of ideal

nutrient water quality with the aim of preserving this state; (2) in situ reference site determinations for

moderately degraded waters; (3) hind casting for historical information from past higher nutrient quality

conditions to determine the reference condition when no reference sites remain; (4) use of loading

estimations from reference quality subestuarine tributary systems and projection to the estuary; and (5)

options for establishing coastal nutrient reference conditions including a Nutrient Criteria Program pilot

demonstration project.

Once the reference condition(s) has been determined, the RTAG then addresses the historical

perspective; considers the need for models to project future consequences; considers the potential effect

on receiving waters; and employs its own good judgment in collectively determining the appropriate

criteria values for each of the variables to protect the waters of concern and their designated uses. A

procedure is also suggested to equate the multiple criteria variables in a comprehensive dimensionless

index score. The manual concludes with a chapter on model development and applications to the criteria

program, and a chapter describing the application and implementation of nutrient criteria with emphasis

on EPA Standards and Monitoring Divisions and a description of a comprehensive ten step sequential

technique for water resource management.

This comprehensive progression from data collection to reference condition determination to criteria

development and management responses, is intended to help users achieve the restoration and protection

of the nutrient water quality of the nation’s estuarine and near-coastal marine water resources.
Nutrient Criteria—Estuarine and Coastal Waters 1-1

Background

Definition of Estuaries and Coastal Systems

Nature of the Nutrient Overenrichment Problem in 

Estuarine and Coastal Marine Waters

CHAPTER 1

Introduction and Objectives

Man has had a long and intimate association with the sea. It has borne his commerce and brought food

to his nets; its tides and storms have shaped the coast where his great cities have grown; the broad

estuaries have provided safe harbors for his ships; and the rhythm of its tides has taught him the

mathematics and science with which he now reaches for the stars (U.S. Department of the Interior 1969).

1.1 BACKGROUND

Nutrient overenrichment is a major cause of water pollution in the United States. The link between

eutrophication—the overenrichment of surface waters with plant nutrients—and public health risks has

long been presumed. However, human health concerns such as (1) Escherichia coli and the spread of

disease in sewage-enriched waters; (2) trihalomethanes in chlorine-treated eutrophic reservoirs; (3) the

incidence of nutrient-stimulated hazardous algal blooms in eutrophic estuarine surface waters with

suspected attendant human illnesses, including recent Pfiesteria investigations; and (4) the relationship of

phytoplankton blooms in nutrient-enriched coastal waters of Bangladesh to cholera outbreaks (Scientific

American, December 1998) all suggest that overenrichment pollution is not only an aesthetic, aquatic

community problem, but also a public health problem.

The purpose of this document is to provide scientifically defensible technical guidance to assist States,

authorized Tribes, and other governmental entities in developing numeric nutrient criteria for estuaries

and coastal waters under the authority of the Clean Water Act (CWA), Section 304a. The objective is to

reduce the anthropogenic component of nutrient overenrichment to levels that restore beneficial uses

(i.e., described as designated uses by the CWA), or to prevent nutrient pollution in the first place. The

primary users of this manual are State/Tribal and Federal agency water quality management specialists

and related interest groups. The manual is intended to facilitate an understanding of cause-and-effect

relationships in these complex systems and serve as a guide for nutrient criteria development, a resource

of technical information, a summary of the scientific literature, and a brief technical account of the

ecological structure and function of estuaries and coastal waters to facilitate an understanding of these

complex systems.

To combat the nutrient enrichment problem and other water quality problems, EPA published the Clean

Water Action Plan, a presidential initiative, in February 1998. Building on this initiative, EPA developed

a report entitled National Strategy for the Development of Regional Nutrient Criteria (U.S. EPA 1998a). 

Criteria form the scientific basis, or yardstick, for ensuring that a desired result will occur because of a

particular form of environmental stress, in this case nutrient overenrichment. The strategic report

outlines a framework for development of waterbody type-specific technical guidance with emphasis on

the reference condition approach that can be used to assess nutrient status and develop region-specific
1-2 Nutrient Criteria—Estuarine and Coastal Waters

numeric nutrient criteria. This technical guidance builds on that strategy and provides guidance for

nutrient criteria development for estuaries and coastal waters. Because estuaries and coastal waters lie at

the interface of the land and include various ecoregions and their rivers, this manual departs somewhat

from the freshwater manuals (e.g., Lakes and Reservoirs, EPA-822-B00-001, and Rivers and Streams,

EPA-822-B-00-002; also available on the EPA web site: www.epa.gov/ost/standards/nutrient.html in

PDF format) and considers both land-based ecoregions and coastal ocean provinces as the geographic

framework. The freshwater nutrient guidance manuals used the ecoregion and subecoregion as the

predominant geographic operational units. 

Because of differing geographic and climatic conditions among the East, Gulf, and West Coasts, uniform

national criteria for estuarine and coastal waters are not appropriate; they should be developed at the

State, regional, or individual waterbody levels. Figures 1-1a,b illustrate the pertinent ecoregions

(including geologic province) of the continental United States associated with coastal and estuarine

waters. In some cases, multiple criteria may be required for large systems with extended physical

gradients. This manual therefore does not provide guidance on how to set nationwide criteria, but

provides State water resource quality managers with guidance on how to set nutrient criteria themselves

relative to EPA regional criteria. This approach is in contrast to toxic chemical criteria, which tend

toward single national numbers with appropriate modifiers (e.g., water hardness for metals). It explores

some approaches to classification of estuaries and coastal shelf systems. The ability to develop useful

classification schemes is still in a highly developmental stage and needs considerable improvement. The

manual describes a minimum set of variables that are recommended for criteria development and

describes methods for developing appropriate values for these criteria. It also provides information on

sampling, monitoring, data processing, modeling, and approaches to implementation and management

responses.

1.2 DEFINITION OF ESTUARIES AND COASTAL SYSTEMS

It is important to have a clear view of the ecosystems that are the focus of this manual. The term

“estuary” has been defined in several ways. For example, a classical definition of estuaries focuses on

selected physical features—e.g., “semi-enclosed coastal waterbodies which have a free connection to the

open sea and within which sea water is measurably diluted with freshwater derived from the land”

(Pritchard 1967) (see Kjerfve 1989 for expanded definition). This definition is limited because it does

not capture the diversity of shallow coastal ecosystems today often lumped under the rubric of estuary. 

For example, one might include tidal rivers, embayments, lagoons, coastal river plumes, and riverdominated coastal indentations that many consider the archetype of estuary. To accommodate the full

range of diversity, the classical definition should be expanded to include the role of tides in mixing,

sporadic freshwater input (e.g., Laguna Madre, TX), coastal mixing near large rivers (e.g., Mississippi

and Columbia Rivers), and tropical and semitropical estuaries where evaporation may influence

circulation. Also, reef-building organisms (e.g., oysters and coral reefs) and wetlands (e.g., coastal

marshes) influence ecological structure and function in important ways, so that biology has a role in the

definition. 
Nutrient Criteria—Estuarine and Coastal Waters 1-3 Figure 1-1a. Draft aggregation of Level III ecoregions for the National Nutrient Strategy illustrating those areas most re

estuarine criteria development.
1-4 Nutrient Criteria—Estuarine and Coastal Waters

Figure 1-1b. Coastal provinces.
Nutrient Criteria—Estuarine and Coastal Waters 1-5

As will be shown, water depth plays a role in the relative importance of sediment-water column fluxes of

materials, including nutrients. These features paint a picture of high ecosystem diversity, where

prediction of susceptibility to nutrient overenrichment is still a scientific challenge and often requires a

great deal of site-specific information. It is because of this diverse response that reference conditions are

a part of nutrient criteria development.

Coastal waters are defined in this manual as those marine systems that lie between the mean highwater

mark of the coastal baseline and the shelf break, or approximately 20 nautical miles offshore when the

continental shelf is extensive. This area will hereafter be referred to as coastal or near-coastal waters. 

Most States have legal jurisdiction out to the 3-nautical-mile limit. However, coastal oceanic processes

beyond this limit may influence nutrient loading and system susceptibility within the 3-mile zone.

1.3 NATURE OF THE NUTRIENT OVERENRICHMENT PROBLEM IN ESTUARINE

AND COASTAL MARINE WATERS

Scope and Magnitude of the Problem

Nutrient overenrichment problems are perhaps the oldest water quality problems created by humankind

(Vollenweider 1992) and have antecedents that extend into biblical history. The basic cause of nutrient

problems in estuaries and nearshore coastal waters is the enrichment of freshwater with nitrogen (N) and

phosphorus (P) on its way to the sea and by direct inputs within tidal systems. Eutrophication, an aspect

of nutrient overenrichment, is portrayed in Figure 1-2. In recent decades, atmospheric deposition of N

has been an important contributing factor in some coastal ecosystems (Vitousek et al. 1997, Paerl and

Whitall 1999). 

In U.S. coastal waters, nutrient overenrichment is a common thread that ties together a diverse suite of

coastal problems such as red tides, fish kills, some marine mammal deaths, outbreaks of shellfish

poisonings, loss of seagrass and bottom shellfish habitats, coral reef destruction, and hypoxia and anoxia

now experienced as the Gulf of Mexico’s “dead zone” (NRC 2000, Rabalais et al. 1991). Additionally,

recent evidence suggests that nutrient enrichment can exacerbate human health effects (Colwell 1996). 

These symptoms of nutrient overenrichment often are preceded by primary symptoms (e.g., an increase in

the rate of organic matter supply, changes in algal dominance, and loss of water clarity) followed by one

or more secondary symptoms listed above (Figure 1-3). Nixon (1995) defined eutrophication as an

increase in the rate of supply of organic matter to a waterbody. In this manual, nutrient overenrichment

is defined as the anthropogenic addition of nutrients, in addition to any natural processes, causing

adverse effects or impairments to beneficial uses of a waterbody. The scientific literature still uses

overenrichment and eutrophication as synonyms. The terms have different meanings, however, because

eutrophication is a natural process in freshwater lakes and presumably in coastal marine waters. An

argument can be made that nutrient stress on coral reefs can cause a loss of symbiotic algae (i.e.,

dinoflagellates), resulting in loss of organic matter and death of the coral colony, a condition not

consistent with eutrophication in the strict sense.
1-6 Nutrient Criteria—Estuarine and Coastal Waters

Figure 1-2. The eutrophication process. Eutrophication occurs when organic matter increases in an ecosystem. 

Eutrophication can lead to hypoxia when decaying organic matter on the seafloor depletes oxygen, and the

replenishment of the oxygen is blocked by stratification. The flux of organic matter to the bottom is fueled by

nutrients carried by riverflow or, possibly, from upwelling that stimulates growth of phytoplankton algae. This flux

consists of dead algal cells together with fecal pellets from grazing zooplankton. Sediment coupled nitrificationdenitrification is shown as well as NO3 transport into sediments when it can be identified. Source: modified from

CENR 2000.

Despite several decades of progress in reducing nutrient pollution from waste treatment facilities,

nutrient runoff from farms and metropolitan areas, often far inland, has gone unabated or actually

increased (The Pew Oceans Commission: www.pewoceans.org; Marine Pollution in the United States:

Significant Accomplishments, Future Challenges, 2001; Mitsch et al. 2001). Interestingly, early marine

scientists considered nutrients as a resource, not a problem (Brandt 1901), and reflected on ways to

fertilize coastal seas to increase biological production. In fact, in the 1890s Brandt concluded that N was

the primary limiting nutrient in marine waters and that nitrification and denitrification were important

processes in the N cycle.

Nutrient overenrichment of estuaries and nearshore coastal waters from human-based causes is now

recognized as a national problem on the basis of CWA 305b reports from coastal States that list waters

whose use or uses are impaired; these figures vary from 25% to 50% of the waters surveyed. The

National Oceanic and Atmospheric Administration’s (NOAA) National Estuarine Eutrophication

Assessment (Bricker et al. 1999) indicated that about 60% of the estuaries out of 138 surveyed exhibited

moderate to serious overenrichment conditions. Nutrient overenrichment of coastal seas now has

international implications (NRC 2000) and is especially well documented for coastal systems of Europe
Nutrient Criteria—Estuarine and Coastal Waters 1-7

Figure 1-3. Expanded nutrient enrichment model. Source: Bricker et al. 1999.

(Justic 1987, Jansson and Dahlberg 1999, Gerlach 1990, cited in Patsch and Radach 1997, Radach 1992),

Australia (McComb and Humphries 1992), and Japan (Okaichi 1997). The problem is likely

underreported for developing nations. Currently, the European Union has initiated an effort to develop

nutrient criteria for surrounding fresh and marine waters (personal communication, U. Claussen, German

Environmental Protection Agency).

In summary, these examples demonstrate that both N and P may limit phytoplankton biomass production

depending on season, location along the salinity gradient, and other factors. Nutrient overenrichment

problems have been present from early history, especially in estuaries downstream of cities, and the

nutrient criteria development approach that follows is a new element in EPA’s effort to address these

longstanding problems.
1-8 Nutrient Criteria—Estuarine and Coastal Waters

1.4 THE NUTRIENT CRITERIA DEVELOPMENT PROCESS

Preliminary Steps

It is impossible to recommend a single national criterion applicable to all estuaries. Natural enrichment

varies throughout the geographic and geological regions of the country, and these subdivisions must be

considered in the development of appropriate nutrient criteria. For example, “drowned river estuaries”

may exhibit a range of inherent or ambient natural enrichment conditions from less than 1.3 μM TP in the

thin soils of the Northeast to 2.6 μM TP in the delta regions of the South and Gulf of Mexico.

Although lakes and reservoirs and streams and rivers may be subdivided by classes, allowing reference

conditions for each class and facilitating cost-effective criteria development for nutrient management, 

except for barrier island estuaries and mangrove bays in a given area this is not feasible for estuaries. A

major distinction between this manual and the one prepared for lakes and reservoirs is that estuarine and

coastal marine waters tend to be far more unique, and development of individual waterbody criteria

rather than for classes of waterbodies (such as glacial temperate lakes) is a greater likelihood. Also,

estuaries will likely require classification by residence time or subdivision by salinity or density

gradients.

Consequently, it will be necessary in many cases to determine the natural ambient background nutrient

condition for each estuary or coastal area so that the eutrophication caused by human development and

abuse can be addressed. Human-caused eutrophication is the focus of this manual, but the development

of nutrient criteria, frequently on a waterbody-specific basis, will require another major distinction for

coastal marine criteria development. In the absence of comparable reference waterbodies, the historical

record of inherent and cultural enrichment may be particularly significant to developing reference

conditions of a particular estuary or coastal reach. The historical perspective is always important to

criteria development, but in this instance it may also be essential to reference condition determination.

An outline of the recommended process for coastal and estuarine criteria development is as follows: (1)

Investigation of historical information to reveal the nutrient quality in the past and to deduce the ambient,

natural nutrient levels associated with a period of lesser cultural eutrophication, (2) determination of

present-day or historical reference conditions for the waterbody segment based on the least affected sites

remaining, such as areas of minimally developed shoreline, of least intrusive use, fed by those tributaries

of least developed watersheds, (3) use of loading and hydrologic models to best understand the density

and flow gradients, including tides, affecting the nutrient concentrations, (4) the best interpretation of

this information by the regional specialists and Regional Technical Assistance Group (RTAG)

responsible for developing the criteria, and (5) consideration of the consequences of any proposed

criteria on the coastal marine waters that ultimately receive these nutrients to ensure that the developed

criteria provide for the attainment and maintenance of these coastal uses. This concept, as illustrated in

Figure 1-4, is the basis for the National Nutrient Criteria Program and is explained throughout this text.

In deriving the reference condition (Figure 1-5), the extreme values of hypereutrophy on one hand and

pristine or presettlement conditions on the other can be estimated from monitoring, historical records, 
Nutrient Criteria—Estuarine and Coastal Waters 1-9

Figure 1-4. Elements of nutrient criteria development and their relationships in

the process.

Figure 1-5. Derivation of the reference condition and the National Nutrient Criteria Program using TP,

TN, and chlorophyll a as example variables. Clarity or Secchi depth would be on a reversed scale. 

Protectivity nutrient criteria should be between pristine conditions and present reference conditions, i.e., the

most “natural” attainable.
1-10 Nutrient Criteria—Estuarine and Coastal Waters

and paleoecological determinations. The reference condition and the derived criteria are scientifically

based estimates expected to be a present-day approximation of the natural state of the waters approaching

but not likely duplicating pristine conditions. They include a conscious decision to use areas of least

human impact as indicators of low cultural eutrophication. A measure of practical judgment is also

necessary where scientific methods and data are not adequate.

The use of minimally impacted reference sites has been adapted from biological criteria development and

is endorsed by EPA’s Science Advisory Board (U.S. EPA 1992). Minimal impacts provide a baseline

that should protect beneficial uses of the Nation’s waters. The term “minimally impacted” implies a high

percentage of conditions in reference locations and a low percentage of conditions in all locations (i.e.,

some enrichment is allowed, but not enough to cause adverse local effects or adverse coastal receiving

water effects). The upper end of the data distribution range from reference sites represents the threshold

of a reference condition, whereas lower percentiles represent high-quality conditions that may not or

cannot be achieved. The upper 25th percentile represents an appropriate margin of safety to add to the

minimum threshold, excludes the effect of spurious outliers, and serves as a sufficiently protective value. 

Where sufficient data are available, comparison and statistical analysis of causal and response variables

can help determine effect thresholds and further refine reference conditions (see Figure 6-2).

Establishing the reference condition is but one element of the criteria development process. Reference

condition values are appropriately modified on the basis of examination of the historical record (most

important), modeling, expert judgment, and consideration of downstream effects.

Strategy for Reducing Human-Based Eutrophication

Six key elements are associated with the strategy for reducing human-based eutrophication (U.S. EPA

1998):

• EPA believes that nutrient criteria need to be established on an individual estuarine or coastal water

system basis and must be appropriate to each waterbody type. They should not consist of a single

set of national numbers or values because there is simply too much natural variation from one part

of the country to another. Similarly, the expression of nutrient enrichment and its measurement

vary from one waterbody type to another. For example, streams do not respond to phosphorus and

nitrogen in the same way that lakes, estuaries or coastal waters.

• Consequently, EPA has prepared guidance for these criteria on a waterbody-type and

region-specific basis. With detailed manuals available for data gathering, criteria development, and

management response, the goal is for States and Tribes to develop criteria to help them deal with

nutrient overenrichment of their waters and protect designated uses.

• To help achieve this goal, the Agency has initiated a system of EPA regional technical and financial

support operations, each led by a Regional Nutrient Coordinator—a specialist responsible for

providing the help and guidance necessary for States or Tribes in his or her region to develop and

adopt criteria. These coordinators are guided and assisted in their duties by a team of inter-Agency
Nutrient Criteria—Estuarine and Coastal Waters 1-11

and intra-Agency specialists from EPA headquarters. This team provides both technical and

financial support to the RTAGs created by these coordinators so the job can be completed and

communication maintained between the policymaking in headquarters and the actual environmental

management in the regions.

• EPA will develop basic ecoregional coastal ocean province nutrient criteria for waterbody types. 

The Regional Teams and States/Tribes can use these values to develop criteria protective of

designated uses; the Agency also may use these values if it elects to promulgate criteria for a State

or Tribe. These criteria, once adopted by States and authorized Tribes into water quality standards,

will have value in two contexts: (1) as decisionmaking benchmarks for management planning and

assessment and (2) as the basis of National Pollution Discharge Elimination System (NPDES)

permit limits and Total Maximum Daily Load (TMDL) target values. The Standards and Health

Protection Division of the EPA Office of Water will be developing implementation guidance for

these latter applications.

• EPA plans to provide sufficient information for States and Tribes to begin adopting nutrient

standards by 2003.

• States/Tribes are expected to monitor and evaluate the effectiveness of nutrient management

programs implemented on the basis of the nutrient criteria. EPA intends the criteria guidance to

reflect the “natural,” minimally impaired condition of a given estuary or coastal water or the class

of these systems, respectively. Once water quality standards are established for nutrients on the

basis of these criteria, the relative success or failure of any management effort, either protection or

remediation, can be evaluated.

Thus, the six elements of the National Nutrient Criteria Program describe a process that encompasses

taking measurements of the collective water resources of an area, establishing nutrient criteria for

evaluating the discrete waters within that region, assessing individual waterbodies against these criteria

and associated standards, designing and implementing the appropriate management, and, finally,

evaluating its relative success.

Nutrient Criteria Development Process

The activities that compose the nutrient criteria development process are listed below in the order

generally followed, and the subsequent chapters of this document follow this sequence. Figure 1-6

presents a schematic illustration of the process with parallel, corresponding chapter headings.
1-12 Nutrient Criteria—Estuarine and Coastal Waters

Figure 1-6. Flowchart of the nutrient criteria development process.
Nutrient Criteria—Estuarine and Coastal Waters 1-13

# Preliminary Steps for Criteria Development (Chapter 1)

Establishment of Regional Technical Assistance Groups

The Regional Nutrient Coordinator in each EPA multistate region should obtain the involvement of key

specialists (e.g., estuarine and marine ecologists, water resource managers, oceanographers, stream and

wetland ecologists, water chemists, and agricultural and land-use specialists) with respect to the

waterbodies of concern. These experts should be recruited from other Federal and State agencies. 

Experts from academia and industry may serve as technical advisors on an as need basis but not official

voting members of the RTAG. 

Particular Federal agencies of interest are the U.S. Geological Survey (USGS); Natural Resources

Conservation Service (NRCS); National Oceanic and Atmospheric Administration (NOAA); National

Marine Fisheries Service (NMFS) and National Ocean Survey (NOS); U.S. Department of the Interior;

National Park Service (NPS); National Seashores; the U.S. Fish and Wildlife Service (USFWS); U.S.

Army Corps of Engineers (USACE); and, in certain areas of the country the Bureau of Land Management

(BLM) or special government agencies such as river basin commissions and inter-State commissions. 

Similarly, for information and education activities, the National Sea Grant Program and for agriculture,

the USDA Cooperative Extension Service are valuable resources.

State agencies with responsibilities relevant to this effort are variously named, but are commonly referred

to as Department of Natural Resources, Department of Water Resources, Department of the

Environment, Department of Environmental Management, Fisheries and Wildlife Management, State

Department of Agriculture, State Department of Forestry, or other land-use management agencies. Most

state land-grant universities have faculty talent important to natural resource and nutrient management,

and almost all colleges and universities have applied science faculty with research interests and talents

appropriate to this initiative.

In selecting participants for the group, diverse expertise is an obvious prerequisite, but willingness to

cooperate in the group effort, integrity, and a lack of a strong alternative interest are also important

factors to consider in selecting these essential people who must make collective and sometimes difficult

determinations.

The experts chosen will constitute the RTAG, which will be responsible for developing more refined

nutrient criteria guidance for their respective estuaries and coastal waters. The RTAG should be large

enough to have the necessary breadth of experience, but small enough to effectively debate and resolve

serious scientific and management issues. A membership of about 30 approaches an unwieldy size,

although that number may initially be necessary to maintain an effective working group of half that size. 

EPA expects that States and authorized Tribes will use the information developed by the RTAGs when

adopting nutrient criteria into their water quality standards. The RTAG is intended to be composed of

scientists and resource managers from Federal agencies and their State counterparts. The RTAG should

not delegate its responsibility with the private sector. The perspectives of private citizens, academicians,

and special interest groups are important, and these and other members of the public may attend RTAG
1-14 Nutrient Criteria—Estuarine and Coastal Waters

meetings and offer opinions when invited, but the final deliberations and decisions are the responsibility

of the Federal and State members of the RTAG—the States when adopting nutrient criteria into their

water quality standards, and the EPA when determining whether to approve or disapprove such criteria. 

They must also be able to meet and debate the issues without undue outside influence.

As a matter of policy, however, EPA encourages the RTAGs to regularly provide access and reports to

the public. The meetings should generally be open to the public and the schedule of those meetings

published in the local newspapers. At a minimum, RTAGs are encouraged to hold regular “stakeholders”

meetings so that environmental, industrial, and other interests may participate via a separate public forum

associated with responding to the group’s efforts. It is important that citizens and public groups be

involved, and any significant determinations of the RTAG should include a public session at which a

current account of activities and determinations is presented and comments acknowledged and

considered. In addition, where specific land uses or practices are addressed, those property owners,

farmers, fishermen, or other involved parties should be consulted in the deliberation and decisionmaking

process.

It is reasonable to expect the RTAG to meet monthly, or at least quarterly, with working assignments and

assessments conducted between these meetings. To coordinate activities among the 10 RTAGS, and with

the National Nutrients Team, regular conference calls are recommended. At these sessions, new

developments in the Program, technical innovations and experiences, budgets, and policy evolutions will

be conveyed and discussed. In the same context, an annual meeting of all Regional Nutrient

Coordinators, State representatives, and involved Federal agencies should be held each spring in or near

Washington, DC. At this meeting, major technical reports are presented by specialists and issues

significant to the Program are discussed.

The composition and coordination discussed above are intended to establish the shortest possible line of

communication between the State, region, and national Program staff members to promote a rapid but

reasoned response to changing issues and techniques affecting nutrient management of our waters. This

format is also designed to be responsive to the water resource user community without becoming a part

of user conflicts.

Delineation of Nutrient Ecoregions/Coastal Province Appropriate to the Development of Criteria

The initial step in this process has been taken through the creation of a national nutrient ecoregion map

consisting of 14 North American subdivisions of the coterminous United States (Figure 1-1). These are

aggregations of Level III ecoregions revised by Omernik (2000). Alaska, Hawaii, and the U.S.

Territories will be subdivided into nutrient ecoregions later, with the advice and assistance of those

States and their governments.

The initial responsibility of each RTAG will be to evaluate the present ecoregional map with respect to

variability on the basis of detailed observations and data available from the States and Tribes in that EPA

region. This preliminary assessment will further depend on the additional nutrient water quality data
Nutrient Criteria—Estuarine and Coastal Waters 1-15

obtained by those States. The databases, especially with respect to selected reference sites, may be used

to refine the initial boundaries of the map in each EPA region.

EPA recognizes that the coastal margins of these ecoregions will be of the greatest concern to the States

developing estuarine and coastal marine criteria, but in some instances watersheds will extend a

considerable distance inland. In any case, the consistent application of the ecoregion concept facilitates

both upstream and inland coordination by the RTAGs and States and integrates the coastal efforts with

rivers, lakes, and streams.

# Scientific Basis (Chapter 2)

Chapter 2 emphasizes the role of physical processes interacting with biological processes in modulating

the expression of nutrient enrichment effects and the potential of inaccurately assessing cause and effects

in developing management plans.

# Physical Classification (Chapter 3)

The next step in evaluating the data is to devise a classification scheme for rationally subdividing the

population of estuarine and coastal marine waters in the State or Tribal territory. Because identification

of overenrichment is the objective of nutrient criteria development, trophic classification per se should be

avoided, as should any classification based on levels of human development. Physical characteristics

independent of most human-caused enrichment sources are far more appropriate. 

However, as stated above, many estuarine and some coastal marine areas will probably require individual

attention and development of reference conditions that are site-specific or at least specific to waterbody

segments. Within these contiguous segments, the reference stations should have similar residence time,

salinity, general water chemistry characteristics, depth, and grain size or bottom type.

Once the waters have been subdivided and classified, it is important to select the key indicator variables

of concern and determine how much information is available on the enrichment status of these stations.

# Selection of Indicator Variables (Chapter 4)

Chapters 4 through 7 describe the variables for which EPA anticipates developing 304 (a) criteria for

nutrients in estuaries and coastal waters and how they should be sampled, preserved, and analyzed. 

Although a wide variety of indicator variables may be possible, this technical manual describes

development of numerical criteria for total phosphorus (TP) and total nitrogen (TN) as primary nutrient

causal variables of eutrophication, and measures of algal biomass (e.g., chlorophyll a for phytoplankton

and ash-free dry weight for macroalgae) and a measure of water clarity (e.g., Secchi depth or electronic

photometers) as primary variables of eutrophic response. In those systems that have hypoxia or anoxia

problems, dissolved oxygen also should be added as a primary response variable. States or Tribes may

elect to include other indicators as well, but the four primary variables and dissolved oxygen as indicated 

are recommended as the essential indicators. Other variables are loss of seagrass/submerged aquatic

vegetation (SAV), benthic macroinfauna, iron, and silica as well as other indicators of primary and

secondary productivity.
1-16 Nutrient Criteria—Estuarine and Coastal Waters

State and Federal agency records are the basis for an initial data search. In many States, water quality

information resides in more than one agency. For example, Maryland has a Department of Natural

Resources and a Department of the Environment, both of which retain water quality records. To

compound the data search problem further, States may also have pertinent data sets in their Department

of Fisheries and Department of Public Health. It is wise to initiate the search for information with calls

and questionnaires to colleagues in the State or Tribal agencies likely to be involved so an appropriate

list of contacts and data sets can be compiled. In doing so, regional Federal agencies should not be

overlooked either. These include the agencies described above in the selection of RTAG members.

# Nutrient Data Collection and Assessment (Chapter 5)

EPA has initiated the data collection and assessment process by screening the existing STORET and

ODES databases for information on lakes, reservoirs, streams, estuaries and coastal waters with respect

to the four initial parameters, and dissolved oxygen where appropriate (see reference to Chapter 4 above).

These primary variables were originally selected for robustness and conservativeness of estimation;

however, the preliminary screening of the STORET data revealed that these measurements are also

relatively abundant in the database. 

Although this is an entirely appropriate starting point for nutrient criteria development, States and Tribes

are not required to confine their investigations and data selection to only these variables. States and

Tribes are encouraged to select additional measures that contribute to the best assessment of the

enrichment of their regional waters and protect designated uses. In particular, it is advisable to use both

causal indicators and response indicators as mentioned above.

Combining nutrient and biological system response information will yield the most definitive and

comprehensive criteria. To use only causal or only response variables in the criteria puts the State or

Tribe in jeopardy of not protecting the designated uses. For example, a highly enriched estuarine system

with a rapid flushing rate may appear to be in attainment when only the biota and dissolved oxygen are

measured, but the load of nutrients being delivered downstream in its coastal discharge plume is

degrading the receiving waters. Using a balanced combination of both causal and response variables in

the criteria, together with careful attention to tidal and seasonal variability, should mitigate against falsepositive or false-negative results.

Chapters 4 and 5 both discuss proper sampling, preservation, and analysis of samples. Seasonality,

spatial distribution of sample sites, composite versus discrete sampling, and fixed station versus stratified

random sampling are also explored.

Establishing an Appropriate Database

Review of Historical Information. Historical information, including sediment core analysis, is important

to establish a perspective on the condition of a given waterbody. Has its condition changed radically in

recent years? Is the system stable over time? What is the variability? Has there been a trend up or down

in trophic condition? Only an assessment of the historical record can provide these answers. Without

this information, the manager risks setting reference conditions and subsequent criteria on the basis of
Nutrient Criteria—Estuarine and Coastal Waters 1-17

present condition alone, which may in fact be a degraded state. Valid historical information places the

current information in its proper perspective and is particularly important to coastal and estuarine

nutrient criteria development because of the difficulty in establishing classes and the scarcity of reference

waterbodies.

Data Screening. The first step in assessing historical or current data is to review the material to

determine its suitability to support nutrient criteria development. Anecdotal information and

observations are valuable, but the sources must be carefully considered. Fishermen’s accounts, local

sport-fishing news stories, and observational logs of scientific field crews are all legitimate sources of

information, but they are subject to different levels of scrutiny before a trend is determined. The same

applies to databases. Nutrient information gathered for identifying failing wastewater treatment plants

cannot be assessed in the same light as similar data collected to determine overall water quality or trophic

state. The analytical procedures used, type of sampling design and equipment, and sample preservation

are other variables that must also be considered in any data review and compilation. Once this screening

is done, the compiled data may be sorted according to station location, physical characteristics, relative

depth, time, and date, and then analyzed for the establishment of reference conditions.

# Establishing Reference Conditions (Chapter 6)

Candidate reference locations can be determined from compiled data with the help of regional experts

familiar with the waters of the area. Classification will be an important first step and should be based on

physical characteristics of the waterbodies, including morphology, geological origin, and hydrologic

factors such as residence time, flow characteristics, tidal processes, and freshwater-saltwater

interchanges. An estuary may then be subclassified into lower, medium, and upper salinity regimes. 

Specialists can also help to select the least culturally impacted sites or stations within each area. 

Three candidate approaches are recommended for development of tidal estuarine reference conditions. 

Two more approaches use loading information within the fluvial watershed. A sixth approach is

described for coastal waters. Where several replicate systems occur, each classified as near-pristine

based on recent data (e.g., past 10 years), then one can apply a frequency distribution approach, and this

manual recommends that the upper 75th percentile be used as a starting point. If some minor nutrient

enrichment is present, then all the data would be considered and, in this case, the lower 25th percentile is

suggested. In the case of significant nutrient-based environmental degradation, where reference sites

cannot be identified from current monitoring data, then hind-casting with ambient data is recommended. 

There are three approaches: (1) empirical in situ data analysis, (2) sediment core or paleoecological

analysis, and (3) model hind-casting. Interpretation of this approach is potentially sensitive to

confounding by physical factors (e.g., freshwater inflows). The watershed approach is load-based. Here,

one attempts to locate a relatively nutrient-unenriched tributary, or stream segment, that is approximately

representative of the watershed, and extrapolate the nutrient load for the entire watershed. This can be

done empirically or, preferably, with models. The coastal approach focuses on changes in the nutrient

regime of estuarine plumes and waters some distance from such plumes. An index approach is described

that accounts for variability and facilitates identification of natural enrichment (e.g., upwelling). Longterm monitoring is required to distinguish anthropogenic effects from natural variability.
1-18 Nutrient Criteria—Estuarine and Coastal Waters

# Criteria Development (Chapter 7)

Nutrient Criteria Components

The move from data review and data gathering to criteria development involves a sequence of five

interrelated elements:

• Examination of the historical record or paleoecological evidence for evidence of a trend.

• Determination of a reference condition using one of several alternative approaches. Remember that

the reference condition, however derived, is only part of the criteria development process.

• Use of empirical modeling or surrogate data sets in some instances where insufficient information

exists. This may be the case especially in estuaries with insufficient hydrological data, or

significantly developed or modified watersheds.

• Objective and comprehensive interpretation of all of this information by a panel of specialists

selected for this purpose (i.e., the RTAG). These experts should have established regional

reputations and expertise in a variety of complementary fields such as oceanography, estuarine

ecology, nutrient chemistry, and water resource and fisheries management.

• Finally, the criterion developed for each variable should reflect the optimal nutrient condition for

the waterbody in the absence of cultural impacts and protect the designated use of that waterbody. 

Second, it must be reviewed to ensure that the proposed level does not entail adverse nutrient

loadings to downstream waterbodies. In designating uses for a waterbody and developing criteria

to protect those uses, the State or Tribe must consider the water quality standards of downstream

waters (40 CFR 131.10 (b)). This concern extends all the way to coastal waters, but in practice the

immediate downstream receiving waters are the area of greatest attention for the resource manager. 

The criteria must provide for the attainment and maintenance of standards in downstream waters. 

A criterion for that estuary or subclass of estuary will not protect downstream water quality

standards, it should be revised accordingly.

Once the initial criteria (either Regional or State/Tribal) have been selected, they can be verified and

calibrated by testing the sampling and analytical methods and criteria values against waterbodies of

known conditions. This ensures that the system operates as expected. This calibration can be

accomplished either by field trials or by use of an existing database of assured quality. This process may

lead to refinements of either the techniques or the criteria.

Criteria are developed for more than one parameter. For example, all reference sites of a given class may

be determined to manifest characteristics of a particular level for TP concentration, TN concentration,

algal biomass,and water clarity. These four measures, and dissolved oxygen as appropriate, become the

basis for criteria appropriate to optimal nutrient quality and the protection of designated uses. The policy

for criteria attainment will be developed by the State or Tribe in consultation with EPA.
Nutrient Criteria—Estuarine and Coastal Waters 1-19

When the estuarine or coastal marine segment in question reveals high TN and TP concentrations, but not

the expected high algal biomass and low water clarity, further investigation is indicated before deciding

whether criteria have been met. Flushing rates, inorganic turbidity, water color, or toxins may be

additional factors influencing the condition of the estuary.

Assessing Attainment With Criteria

An action level then is established for the nutrient criteria that have been selected for each indicator

variable. The list includes two causal variables (TN and TP) and three primary response variables (e.g.,

when dissolved oxygen problems occur this will add an additional variable to the response variables.

Failure to meet either of the causal criteria should be sufficient to prompt action. However, if the causal

criteria are met, but some combination of response criteria are not met, there should be some form of

decision making protocol to resolve the question of whether the waters in question meet the nutrient

criteria. There are two approaches to this:

• Establish a decisionmaking rule equating all of the criteria such as the frequency and duration of

exceedences and the critical combination of response variables requisite for action

• Establish an index that accomplishes the same result by inserting the data into an equation that

relates the multiple variables in a nondimensional comprehensive score

# Management Response (Chapter 8)

There are a variety of possible management responses to the overenrichment problem identified by

nutrient criteria. Chapter 8 describes some regulatory and nonregulatory processes that involve the

application of nutrient criteria. It also presents a 10-step process that allows the resource manager to use

these approaches to improve water resource condition. The emphasis is on developing a scientifically

responsible, practical, and cost-effective management plan.

The chapter also describes three basic categories that encompass all management activities: education,

funding, and regulation. It closes with the admonition to always carefully evaluate the success of the

management project, report results, and continue monitoring the status of the water resource.

# Model Applications (Chapter 9)

A variety of empirical and theoretical models are described and discussed, and two specific illustrations

of the application of models to estuarine nutrient management are presented.

# Appendices

A number of appendices supplement the primary text.

It should be noted that completion of each step may not be required of all water quality managers. Many

State or Tribal water quality agencies may have already completed the identification of designated uses,

classified their estuaries and coastal waters, or established monitoring programs and/or databases for

their programs and therefore can bypass those steps. This manual is meant to be comprehensive in the
1-20 Nutrient Criteria—Estuarine and Coastal Waters

sense that all of the criteria development steps are described; however, the process can be adapted to suit

existing water quality programs.

In any event, a responsible nutrient management plan should meet three conditions. First, the plan and its

component elements must be scientifically defensible; otherwise it might lead to well-intentioned

management actions that are unnecessary or harmful. This is like the admonition to physicians, “above

all do no harm.” Second, effective nutrient management must strive to be economically feasible. The

public and local interests are more likely to support approaches that provide meaningful benefit

compared with their cost. Finally, these approaches should be practical and acceptable to the

communities involved. The approaches should address appropriate social and political issues, such as

conflicts that might exist between public agencies and landowners, agricultural or other resource users, or

between commercial fishermen and recreationists and environmental or industrial groups. Any

management plan may fail if these three general elements are not sufficiently addressed, and it is almost

certain to fail if they are all ignored.
Nutrient Criteria—Estuarine and Coastal Waters 2-1

Controlling the Right Nutrients

Physical Processes, Salinity, Algal Net Primary Production

Nutrient Loads and Concentrations: Interpretation of Effects

Physical-Chemical Processes and Dissolved Oxygen Deficiency

Nutrient Overenrichment Effects and Important Biological

 Resources

Concluding Statement Regarding Nitrogen and Phosphorus

CHAPTER 2

Scientific Basis for

Estuarine and Coastal

Waters Quantitative

Nutrient Criteria

2.1 INTRODUCTION

At the turn of the last century nitrogen and phosphorus were prized as the fuel that fed the great engine

of marine production. Today they are seen as lethal pollutants leading to toxic blooms and suffocation. 

Just as weeds are fine plants growing in the wrong place, nitrogen and phosphorus are essential

chemicals that can get into the wrong places at the wrong times. We should not lose sight of their

critical role in sustaining production (Nixon 2000).

Purpose and Overview

This chapter describes the scientific basis for development of nutrient criteria for estuarine and coastal

waters. A number of scientific issues are addressed to develop nutrient criteria. Water quality managers

can improve their application of science to nutrient criteria development if they consider these systems’

large latitudinal and climatic range, high ecosystem-based variability, complexity, diversity, and broad

range in land-sea margin human activities. These features suggest a high degree of system individuality,

especially at larger scales. These features occur because estuaries and coastal waters are transitional

ecosystems buffeted by variable landward-based freshwater input volumes and constituents, influences of

oceanic provinces, and human disturbances, including nutrient enrichment, superimposed on these

natural regimes (Figure 2-1). Even in a relatively narrow section of coastline, the ecosystem diversity

and variability may be quite large. These characteristics challenge the investigator to develop useful

predictive schemes. Some progress has been achieved, but areas of important uncertainties are also

noted.

Coastal areas, including estuaries and upwelling regions, account for only 10% of the ocean by area but

at least 25% of the ocean’s primary productivity and upwards of 95% of the world’s estimated fishery

yield (Walsh 1988). These areas are also an important organic carbon sink of atmospheric CO2. In

addition, coastal counties account for only 17% of the U.S. landmass, but their population exceeds 141

million. Thus, more than half of the Nation’s population lives in less than one-fifth of the total area, and

this trend is expected to grow (NRC 2000). These statistics underpin the fact that estuarine and open

coastal areas have, and continue to show, stress from human activities including nutrient pollution, as

noted in Chapter 1. These demographics argue strongly for a scientific understanding of how nutrients

flux through estuarine and nearshore coastal ecosystems and impair water quality use. 
2-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-1. Idealized scheme defining the coastal ocean and the coastal zone, with some key biochemical fluxes

linking land and sea and pelagic and benthic processes. The latter are not to scale. Source: Alongi 1998.

Some Important Nutrient-Related Scientific Issues

A large number of issues with a scientific component may complicate nutrient criteria development in

estuaries and open coastal waters. Some of the more important issues are summarized below and are

discussed in more detail later in this and following chapters. These issues illustrate how science

underpins nutrient criteria development.

Determination of which nutrients are causing the problem is critical. In some cases, this will be known

with considerable assurance, but in others further study is advisable. Without such knowledge, it is

difficult to develop reliable nutrient criteria. It is important to understand at what scale one is discussing

the question of nutrient limitation. The term “nutrient limitation” is often used quite loosely and without

formal definition (Howarth 1988). For phytoplankton, Howarth makes the following points and argues

that it matters a great deal which of the following questions is being addressed:

• Limitation of the growth rate of phytoplankton populations currently in a waterbody

• Limitation of the potential rate of net primary production, allowing for possible shifts in the

composition of phytoplankton species

• Limitation of net ecosystem production
Nutrient Criteria—Estuarine and Coastal Waters 2-3

Each of these definitions can be considered “correct,” but each addresses different questions. Clearly,

phytoplankton growing in an oligotrophic environment may be adapted to maximize growth rates under

low nutrient conditions, as evidenced by their organic nutrient composition approaching the Redfield

atomic ratio of C:N:P of 106:16:1 (Redfield 1958; Goldman et al. 1979). An increase in nutrient supply

would likely shift species composition to those adapted to the higher nutrient regime, and net primary

production would potentially increase. Thus, it is plausible that potential net primary production can be

nutrient limited even if the growth rate of currently dominant phytoplankton species is not. If a nutrient

is added to a system and net primary production increases, the system is considered to have been nutrient

limited regardless of whether the species composition has shifted. Similarly, when a nutrient criterion is

exceeded, enrichment is presumed to be of concern even if the system’s productivity has not responded. 

This is the definition used in this manual for addressing effects of nutrient overenrichment.

Why not use net ecosystem production as the preferred definition, as the ecosystem is the level of system

organization that might seem most relevant? For example, the ecosystem was the level of the whole-lake

experiments that contributed to defining P as the primary limiting nutrient for north temperate freshwater

lakes (Schindler 1977). Net ecosystem production equals gross primary production in excess of total

ecosystem respiration. For the biomass of an isolated ecosystem to be maintained, the net ecosystem

organic production must equal or slightly exceed 0. Imports of organic matter can augment the internal

net production. Howarth argues that it is difficult to relate nutrient supplies to net ecosystem production

because the respiration term is sensitive to allochthonous input of organic matter as well as internal net

production. So, for practical reasons, net primary production, which is directly related to algal biomass

production, is the preferred measure of nutrient limitation. 

The import of organic matter, especially in estuaries, can lead to water quality problems (e.g., hypoxia). 

Organic matter input from sewage was historically a major source of organic carbon that drove aquatic

systems toward dissolved oxygen (DO) deficiency through direct microbial heterotrophic activity

(Capper 1983). However, the input of nutrients, whether in organic form followed by recycling or

inorganic form with direct nutrient uptake, is what stimulates potential phytoplankton biomass

production, and this organic matter may contribute to symptoms of nutrient overenrichment identified in

Chapter 1.

It is frequently difficult to distinguish natural ecosystem variability associated with net primary

production from that induced by anthropogenic stress, especially nutrient enrichment, which often is a

consequence of variability in physical processes. An example is the difficulty, even with a 50-year

record, in distinguishing the effects of freshwater flow of the Susquehanna River and co-linear effects of

nutrient loading on Chesapeake Bay phytoplankton biomass production indicated by chlorophyll a (chl a)

concentrations (Harding and Perry 1997). Such indeterminancy is a condition that water quality

managers must contend with, and argues for broad scientific input.

It is important to understand nutrient load and ecological response relationships because of the need to

conduct load allocations (e.g., total maximum daily loads, TMDLs), and it may be necessary to perform

some management triage when systems are poised along a gradient of risk and there are too many
2-4 Nutrient Criteria—Estuarine and Coastal Waters

systems to treat in a timely fashion. Also, as explained later, ecological responses to nutrient enrichment

may be quantitatively related to nutrient load rather than complexity in physical transport and mixing. 

The relationship between N load and seagrass recovery in Tampa Bay, FL, is an example of where

nutrient load was predictive but concentration of N was not (Greening et al. 1997).

As discussed in Chapter 3, classification of estuaries and coastal shelf systems at large scale (e.g.,

Chesapeake Bay versus Delaware Bay) is in an early state of development with regard to predicting many

nutrient enrichment effects. This is because of the relatively high degree of ecosystem individuality at

the larger scale, where comparability among systems tends to breaks down. The result is that scientific

generalizations are usually circumscribed with consequences that may lead to higher management costs. 

Resource managers and environmental scientists should work together to improve predictability of

nutrient enrichment effects because there are too many systems in the Nation to study all estuaries and

coastal systems comprehensively.

These ecosystems exhibit a notable degree of process asymmetry and lag in responses, which means that

a stress at one location and time may show up as a response at another location and time. Additionally,

different mechanisms may result in a similar response (Malone et al. 1999). This type of behavior

enhances the tendency to confound cause-and-effect relationships.

Along the same lines, conceptual models for estuaries (and coastal waters) in particular are still evolving. 

These models suggest that systems modulate stresses so that a single stress does not necessarily result in

a single response (Cloern 2001) (Figure 2-2). This fact alone contributes to ecological uncertainty in

load-response relationships. Conceptual models help define expectations of cause-and-effect

relationships and degree of nutrient-caused impairment, and refine hypotheses. Conceptual models

should be a standard tool for water quality managers.

Antecedent conditions are important. This can be understood in terms of whether enough factors are

present at the right place and time to lead to an integrated response, such as a dinoflagellate bloom. Such

conditions resemble nonlinear dynamics, which may be a major constraint to prediction of effects. Also,

estuaries and nearshore coastal waters are subject to episodic events, which injects considerable

uncertainty into predictions (e.g., Tropical Storm Agnes impacted Chesapeake Bay in June 1972: Davis

and Laird 1976). A relatively large database is often required to determine when effects of such major

events have reached a new steady state.

Estuaries and nearshore coastal waters naturally vary in the type, abundance, and geographical coverage

of biological communities at risk to nutrient overenrichment, largely because of habitat differences. This

variability is partially offset by salinity, which tends to “normalize” biotic community distributions

(Kinne 1964). When ambient historical data are unavailable or sediment cores are ineffective in 
Nutrient Criteria—Estuarine and Coastal Waters 2-5

Figure 2-2. Schematic representation of the contemporary (Phase II) conceptual model

of coastal eutrophication. Advances in recent decades include explicit recognition of (1)

a complex suite of both direct and indirect responses to change in nutrient inputs; (2)

system attributes that act as a filter to modulate these responses; and (3) the possibility of

ecosystem rehabilitation through appropriate management actions to reduce nutrient

inputs to sensitive coastal ecosystems. Source: Cloern 2001.

characterizing resources lost through nutrient overenrichment, it is often difficult to establish an accurate

historical reference or determine the potential recovery from nutrient stress. Apparently, many estuaries

became moderately to highly enriched before effective monitoring programs provided accurate

descriptive information on biotic community distributions and abundance. When all else fails,

professional judgment should be used to estimate reference conditions.

Finally, water quality managers should anticipate that nutrient enrichment will act with other stressors

and forms of ecosystem disturbance and modify their respective ecological expressions (Breitburg et al.

1999).

These considerations suggest that water quality managers may face a large array of uncertainties

regarding nutrient criteria development and implementation for estuaries and nearshore coastal waters. 

This manual attempts to guide application of established scientific principles and to reveal important

uncertainties that bear on nutrient criteria development. This chapter begins with a contextual discussion

of the watershed perspective characterized as the “river-to-ocean continuum.”

River-to-Ocean Continuum: Watershed/Nearshore Coastal Management Framework

This section describes the physical relationship of estuaries and nearshore coastal waters to their

respective water and sedimentary boundaries. This description provides a context for understanding

problems of nutrient overenrichment in coastal ecosystems. Estuaries and nearshore coastal systems

share some features, but important differences reflect how nutrients cause problems.
2-6 Nutrient Criteria—Estuarine and Coastal Waters

Some Important Identifying Features of Estuaries (adapted partly from Cloern 1996)

1. Estuaries are located between freshwater ecosystems (lakes, rivers, and streams; freshwater and

coastal wetlands; and groundwater systems) and coastal shelf systems (Figure 2-2). These

ecological boundary conditions create a transition between contrasting freshwater and open-ocean

ecosystems.

2. Estuaries are relatively shallow; often, on average, only a few meters to a few tens of meters deep. 

This promotes a strong benthic-pelagic coupling that influences nutrient cycling through changes in

system nutrient stoichiometry. A well-developed benthic community participates in nutrient

cycling.

3. River-influenced estuaries are quite different from systems. Vertical mixing is regulated primarily

by the seasonal cycle of heat input and thermal stratification that retards vertical mixing. However,

in estuaries vertical mixing is regulated by a larger and more variable source of buoyancy: the

riverine input of freshwater that acts to stabilize the water column. Also, freshwater input

establishes longitudinal and vertical salinity gradients and drives nontidal gravitational circulation,

a major contributor to flushing. 

4. Estuaries are particle-rich relative to coastal systems and have physical mechanisms that tend to

retain particles. These suspended particles mediate a number of activities (e.g., absorbing and

scattering light, or absorbing hydroscopic materials such as phosphate and toxic contaminants). 

New particles enter with river flow and may be resuspended from the bottom by tidal currents and

wind-wave activity.

5. Many estuaries are naturally nutrient-rich because of inputs from the land surface and geochemical

and biological processes that act as “filters” to retain nutrients within estuaries (Kennedy 1984).

Variability in freshwater discharge is reflected in the estuarine salinity gradient, which has important

consequences for stenohaline organisms, especially nonmotile forms. The salinity gradient of estuaries

has been classified by on the Venice System, and salinity classes approximate the distribution of many

estuarine organisms (Figure 2-3). Changes in salinity (e.g., wet and dry decadal periods) often modify

population distributions and biotic community structure (Carriker 1967). Rivers and lakes process

nutrients and modify nutrient ecological stoichiometry before the material arrives downstream, where

receiving coastal waters further nutrient cycling (Billen et al. 1991). Nutrient cycling occurs along the

continuum; phytoplankton and other algae are key agents of biochemical change (Redfield 1963) (Figure

2-4). Redfield et al. (1958) demonstrated that phytoplankton in active growth phase tend to maintain a

C:N:P ratio close to 106:16:1. Annual rates of net primary production in coastal shelf environments tend

to overlap rates of estuaries, but coastal shelves on average are somewhat lower in magnitude, except in

upwelling areas where rates may, on average, exceed those of estuaries by a factor of two to three (Walsh

1988) (Table 2-1). 
Nutrient Criteria—Estuarine and Coastal Waters 2-7

Figure 2-3. Salinity zones. The Venice System is a well-accepted method of characterizing salinity zones and

covers the salinity ranges from riverine regions to the ocean. The freshwater category in the Venice System has been

modified in this atlas to account for the tidal and nontidal regions found in rivers with estuarine portions. Source: 

Lippson et al. 1979, Environmental Atlas of the Potomac Estuary, MD Department of Natural Resources.

Figure 2-4. Schematic illustrating the central role of phytoplankton as agents of biogeochemical change in shallow

coastal ecosystems. Phytoplankton assimilate reactive inorganic substances and incorporate these into particulate

(POM) and dissolved organic matter (DOM) which support the production of pelagic and benthic heterotrophs. 

Arrows indicate some of the material fluxes between these different compartments. Denitrification has been added to

the figure. Source: Cloern 1996.
2-8 Nutrient Criteria—Estuarine and Coastal Waters

Table 2-1. Categorization of the world’s continental shelves based on location, major river, 

and primary productivity

Latitude

(°)

Region Major

River

Primary Production

(g Cm-2 yr1)

Eastern Boundary Current

0-30 Ecuador-Chile G 1000-2000

Southwest Africa G 1000-2000

Northwest Africa G 200-500

Baja California G 600

Somali coast Juba 175

Arabian Sea Indus 200

30-60 California-Washington Columbia 150-200

Portugal-Morocco Tagus 60-290

Western Boundary Currents

0-30 Brazil Amazon 90

Gulf of Guinea Congo 130

Oman/Persian Gulfs Tigris 80

Bay of Bengal Ganges 110

Andaman Sea Irrawaddy 50

Java/Banda Seas Brantas 110

Timor Sea Fitzroy 100

Coral Sea Fly 20-175

Arafura Sea Mitchell 150

Red Sea Awash 35

Mozambique Channel Zambesi 100-150

South China Sea Mekong 215-317

Caribbean Sea Orinoco 66-139

Central America Magdalena 180

West Florida shelf Appalachicola 30

South Atlantic Bight Altamaha 130-350

Mesotrophic Systems

30-60 Australian Bight Murray 50-70

New Zealand Waikato 115
Table 2-1. Categorization of the world’s continental shelves based on location, major river, and primary 

 productivity (continued)

Latitude

(°)

Region Major

River

Primary Production

(g Cm-2 yr1)

Nutrient Criteria—Estuarine and Coastal Waters 2-9

Argentina-Uraguay Parana 70

Southern Chile Valdivia 90

Southern Mediterranean Nile 30-45

Gulf of Alaska Fraser 50

Nova Scotia-Maine St. Lawrence 130

Labrador Sea Churchill 24-100

Okhotsk Sea Amur 75

Bering Sea Kuskokwim 170

Phototrophic Systems

60-90 Beaufort Sea Mackenzie 10-20

Chukchi Sea Yukon 40-180

East Siberian Sea Kolyma 70

Laptev Sea Lena 70

Kara Sea Ob 70

Barents Sea Pechora 25-96

Greenland-Norwegian Seas Tjorsa 40-60

Weddell-Ross Seas G 12-86

Eutrophic Systems

30-60 Mid-Atlantic Bight Hudson 300-380

Baltic Sea Vistula 75-150

East China Sea Yangtze 170

Sea of Japan Ishikari 100-200

North-Irish Sea Rhine 100-250

Adriatic Sea Po 68-85

Caspian Sea Volga 100

Black Sea Danube 50-150

Bay of Biscay Loire 120

Louisiana/Texas shelf Mississippi 100

Source: Adapted from Walsh, with additional data from Alongi, and Postma and Zijlistra.
2-10 Nutrient Criteria—Estuarine and Coastal Waters

Some Identifying Features of Nearshore Coastal Waters

1. Nearshore coastal waters extend from the coastal baseline at high tide and across the mouths of

estuaries to approximately three nautical miles. Coastal waters are relatively deep compared to

estuaries with depths ranging from a few meters to several hundred meters, depending on coastal

location. 

2. Coastal longshore currents are a principal mechanism to exchange water masses.

3. Upwelling of nutrients from the deep ocean can be locally important.

4. Nearshore coastal systems tend to be particle-rich compared to the open ocean, but much less so

than adjoining estuaries.

5. Nearshore coastal systems have a weaker benthic-pelagic coupling than estuaries mainly because

they are deeper.

Coastal environments in the continental United States show only modest levels of upwelling compared to

well-known upwelling areas, such as coastal Ecuador-Chile. The Gulf Stream, which flows

northeastward along the South Atlantic coast from the Florida Straits to North Carolina, lies close enough

to the shoreline to affect water temperature and circulation of nearshore waters. Dynamic core rings that

slide off to the mainland side of the Gulf Stream affect local conditions. The coastal environment is

dynamic in terms of phytoplankton bloom formation and dissipation (Walsh 1988). This has relevance to

characterization of reference conditions and monitoring for nutrient criteria performance because the

systems, though not as physically dynamic at short temporal scales as estuaries, are still difficult to assess

in terms of average conditions. Synoptic survey tools such as aerial surveillance with fixed-wing aircraft

and satellites can provide wide coverage, including short-term phytoplankton dynamics.

2.2 CONTROLLING THE RIGHT NUTRIENTS

Overview

Chapter 1 introduced the geographical extent and magnitude of the overenrichment problem and

suggested the importance of nitrogen (N) versus phosphorus (P) as limiting nutrients. Several recent

review papers (Downing 1997, Smith 1998, Smith et al. 1999, Conley 2000) and the NRC (2000) volume

concluded that the major nutrients causing overenrichment problems (e.g., algal blooms) in estuaries and

nearshore coastal waters are N and P. Silica (Si) may limit diatom production at relatively high levels of

N and P. Iron is a co-limiting nutrient in some ocean areas and may exert some limitation in shelf

waters,but its importance in open coastal waters usually is secondary to N (NRC 2000). Additionally, P

limits primary production in some tropical nearshore habitats, although study of these systems is limited

(Howarth et al. 1995). Often the addition of both N and P will elicit greater phytoplankton biomass

stimulation than the sum of both nutrients added separately (Fisher et al. 1992). There are reported cases

where both N and P are required to elicit a phytoplankton biomass production response in estuaries

(Flemer et al. 1998), suggesting that N and P supply rates were equally limiting. Tropical lagoons, with
Nutrient Criteria—Estuarine and Coastal Waters 2-11

carbonate sands low in P and unaffected by human activity, also are prone to P limitation. For example,

the seagrass Thalassia testudinum was P-limited in Florida Bay (Powell et al. 1989, Fourqurean et al.

1992a,b).

Tidal fresh and brackish waters in many estuaries typically are more light limited than higher saline

waters (Flemer 1970, Sin et al. 1999). As freshwater fluxes seaward, processes operate to modify

nutrient stoichiometry (e.g., sedimentation of P-absorbed particles, denitrification, and differential

microbial decomposition). A number of temperate estuaries exhibit seasonal shifts in nutrient limitation

with winter-spring P limitation and summer-fall N limitation (D’Elia et al. 1986; Fisher et al. 1992,

Malone et al. 1996) (Table 2-2). The Redfield ratio (C:N:P) of marine benthic plants approximates

550:30:1, substantially richer in organic carbon, much of which is structural material, and indicates that

these plants require less N and P than do phytoplankton (Atkinson and Smith 1982). In summary, the

foregoing results suggest that both N and P criteria are needed, depending on season and local ecosystem

conditions (Conley 2000).

Some Empirical Evidence for N Limitation of Net Primary Production

Three case studies provide some of the strongest evidence available that water quality mangers should

focus on N for criteria development and environmental control (see NRC 2000 for details). One study

involves work in large mesocosms by the University of Rhode Island (Marine Ecosystem Research

Laboratory–MERL) on the shore of Narragansett Bay. Experiments showed that P addition was not

stimulatory, but N or N+P caused large increases in the rate of net primary production and phytoplankton

standing crops (Oviatt et al. 1995). 

In another study, nutrient releases from a sewage treatment plant were monitored in the Himmerfjarden

Estuary south of Stockholm, Sweden, on the Baltic Sea (Elmgren and Larsson 1997). Throughout a 17-

year field experiment (i.e., whole-ecosystem study), the concentration of total N tended to reflect the N

input from the sewage treatment plant, and both abundances of phytoplankton and water clarity were

clearly related to the total N concentration and not to total P. This experiment involved independent

increases and decreases in N and P over the observation period.

A third whole-ecosystem study involved long-term changes in Laholm Bay, Sweden (Rosenberg et al.

1990). Early signs of overenrichment appeared in the 1950s and 1960s and steadily increased over time

(Figure 2-5). Among the earliest reported signs were changes in the composition of macroalgal species. 

Over time the filamentous algae typical of enriched conditions became more prevalent, and harmful algal

blooms (HABs) became more common during the 1980s. These changes correlated best with changes

over the decades in N loads rather than P loads. These field studies are excellent examples of the power

of long-term monitoring of nutrient and biological variables in estuaries (Wolfe et al. 1987). Importantly,

these three ecosystem experiments correlated well with short-term bioassay experiments and ratios of

dissolved inorganic N:P ratios in these ecosystems (NRC 2000). The above whole-system field

experiments and the large preponderance of bioassay data in estuaries and nearshore coastal systems

(Howarth 1988) and generally low inorganic N:P atomic ratios at peak primary production (Boynton et

al. 1982) make a strong case for the widespread importance of N as a controlling nutrient for net coastal
2-12 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-5. Transport of nutrients to Laholm Bay, Sweden. Periods of significant

changes in the marine biota are also indicated (modified from Rosenberg et al. 1990). 

Source: NRC 2000.

Table 2-2. Estuaries exhibiting seasonal shifts in nutrient limitation with 

spring P limitation and summer N limitation

Estuary Reference

Baltic Sea

 Himmerfjarden Estuary, Sweden Graneli et al. 1990, Elmgren & Larsson 1997

 Gulf of Riga, Latvia Maestrini et al. 1997

Roskilde Fjord, Denmark Pedersen & Borum 1996

Bay of Brest, Francea Del Amo et al. 1997

Chesapeake Bay, USAa

 Mainstem Malone et al. 1996

 Patuxent River Estuary D’Elia et al. 1986

 York River Estuary Webb 1988

 Rhode River Estuary Gallegos & Jordan 1997

Delaware Estuary, USA Pennock & Sharp 1994

Neuse River Estuary, USA Mallin & Paerl 1994

a

Systems displaying seasonal dissolved silicate limitation.

 Source: Conley 2000.
Nutrient Criteria—Estuarine and Coastal Waters 2-13

marine primary production and a major contributor to water quality problems. Interpretation of nutrient 

ratios was initially applied in the open ocean by Redfield (1934) and further elaborated on by Redfield

(1958) and Redfield et al. (1963). Boynton et al. suggested that when inorganic N:P ratios for a variety

of estuarine systems are interpreted, atomic ratios less than 10 indicated N limitation and ratios greater

than 20 indicated P limitation (Figure 2-6). Some have suggested that it matters whether the inorganic N

is in the form of ammonium- or nitrate-N. High concentrations of ammonia-N may inhibit nitrate-N

uptake; however, Dortch (1990) reported that this phenomenon is more variable than widely believed. 

Figure 2-7 summarizes major factors that determine whether N or P is more limiting in aquatic

ecosystems where one of these macronutrients is limiting net primary production.

Some Threshold Responses to Nitrogen Overenrichment

Kelly (in press) summarized several generalizations that appear to hold for N overenrichment in

estuaries. Over a range of average dissolved inorganic nitrogen (DIN) from <1 to >20 μM, chlorophyll a

tends to increase at slightly less than 1 μg/L with every 1 μM increase in DIN or approximately about

0.75 μg chl/μM DIN (e.g., see Figure 3-2b in Chapter 3). Evidence is especially strong that N

concentrations can reduce or eliminate growth of estuarine submerged aquatic vegetation (SAV) and

higher salinity seagrasses (Sand-Jensen and Borum 1991; Dennison et al. 1993; Duarte 1995) by both

water column shading and epiphytic overgrowth. Estuarine SAV and seagrasses tend to show light

limitation when surface insolation approximates 11% at the surface of the canopy, but this figure varies

between about 5% and 20% depending on species. Stevenson et al. (1993) transplanted plugs of Ruppia

maritima, Potamogeton perfoliatus, and P. pectinatus in different areas of the Choptank Estuary,

Chesapeake Bay, and reported that survival thresholds occurred when total suspended solids were

between ~15 and 20 mg/L, chlorophyll a was 15 μg/L, DIN was below 10 μM, and PO4 was below 0.35 

μM. Kelly (in press) reviewed a number of studies and suggested that an approximate threshold for

hypoxia occurred at about 80 μM TN (Table 2-3) (normalized TN loading for residence time expressed

in years and divided by depth). These relationships document the importance of N as a major cause of

estuarine water quality impairment. Also, these ecological response thresholds are a useful rule of

thumb, but some deviations are to be expected. In data-poor estuaries, such thresholds are a first-order

target until more adequate data can be developed to establish reference conditions.

Although overenrichment from N causes many symptoms of marine water quality impairment, it is the

interaction of biogeochemical, biological, and physical processes that modulate the effects of a particular

N supply (Cloern 2001) (Appendix A). These relationships had their genesis in the late 19th and early

20th Centuries in northern Europe, especially in German and Scandinavian marine research institutes

(Mills 1989). Water quality managers who understand this interplay will assess cause-and-effect

relationships with a deeper insight. Knowledge of algal nutrient physiology is necessary information, but

it alone is insufficient to explain why blooms occur.

Effects of Physical Forcing on Net Primary Production

Each physical forcing (e.g., river inflows, wind velocity, irradiance, water temperature, and tidal

currents) contributes to phytoplankton population variability by influencing rates of vertical mixing,
2-14 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-6. Summary of nitrogen:phosphorus ratios in 28 sample estuarine ecosystems. 

Horizontal bars indicate the annual ranges in nitrogen:phosphorus ratios; solid triangles

represent the ratio at the time of maximum productivity. Vertical bands represent the

typical range of algal composition ratios (modified from Boynton et al. 1982). Source: 

NRC 2000.
Nutrient Criteria—Estuarine and Coastal Waters 2-15

Figure 2-7. Factors that determine whether nitrogen or phosphorus is more limiting in aquatic

ecosystems, where one of these macronutrients is limiting to net primary production. 

Phytoplankton use nitrogen and phosphorus in the approximate molar ratio of 16:1. The ratio of

available nitrogen in the water column is affected by: (1) ratio of nitrogen:phosphorus in external

inputs to the ecosystem; (2) relative rates of recycling of nitrogen and phosphorus in the water

column, with organic phosphorus usually cycling faster than organic nitrogen; (3) differential

sedimentation of nitrogen in more oligotrophic systems; (4) preferential return of nitrogen or

phosphorus from sediments to the water column due to processes such as denitrification and

phosphorus adsorption and precipitation; and (5) nitrogen fixation (modified from Howarth 1988;

Howarth et al. 1995). Source: NRC 2000. 
2-16 Nutrient Criteria—Estuarine and Coastal Waters

Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL mesocosm

enrichment experiment (source: Kelly in press)

System Area

Depth

Avg

(m)

Annual TN

loading

(mmol m-2)

Res. Time

(mo)

DO

Statusa

Vertical

Mixing

Status

Normalized

TN Loading

(μM)b

Primary

Production 

(g Cm-2 y-1)

Experimentalc(m2)

MERL-control 2.63 5 800 0.9 OK mixed 12 190 (100)

MERL-1X 2.63 5 1,750 0.9 OK mixed 26 270 (115)

MERL-2X 2.63 5 2,950 0.9 OK mixed 44 305 (243)

MERL-4X 2.63 5 4,850 0.9 OK mixed 72 515 (305)

MERL-8X 2.63 5 9,000 0.9 ~H mixed 133 420 (171)

MERL-16X 2.63 5 18,500 0.9 H mixed 274 900 (601)

MERL-32X 2.63 5 34,000 0.9 A mixed 503 1150 (901)

Fieldd

(km2)

Baltic Sea e374,600 55 217 250 H/A stratified 81 ~149-170

Scheldt 277 11.2 13,400 3 H/A ?? 295 ?

Chesapeake Bay f,g 11,542 6 938 7.6 A stratified 98

~380 to 520 

(361-858)

Potomac River f1,210 5.9 2,095 5 H/A stratified 146 ~290 to 325

Guadalupe estuary h551 1.4 548 10 ? ?? 322 ?

551 1.4 2,058 1 ? ?? 121 ?

Ochlocknee Bay 24 1 5,995 0.1 OK 49 ?

Delaware Bay 1,989 9.7 1,900 4 OK stratified 64 ~200 to 400

Narragansett Bay i328 8.3 1,960 0.9 OK weak strat 17 270 to 290

 Providence River j24.13 3.7 13,600 0.083 H stratified 25 ?

 Providence Riv.j,k 24.13 3.7 13,600 0.233 H stratified 70 ?

Boston Harbor l103 5.5 21,600 0.266 ~H weak strat 86 ?

 N. Outer Harbor m13 10 107,692 0.03 OK mixed 27 263 to 546

N. Gulf of Mexico n20,000 30 6500 6 O H/A stratified 107 ~290 to 320

aH= hypoxia, A= anoxia.

bVolumetric TN loading is normalized for residence time to yield an “expected” or potential concentration. The value is calculated as: Annual

TN Loading * Residence time (expressed in years) divided by Depth. Units are thus mmol/m3, or :M. See Kelly 1997a,b; 1998. The value is

not decremented for denitrification or burial, removal processes that have greater effect on concentrations in longer residence time systems (cf.

Nixon et al. 1996, Kelly 1998).

cSee Nixon et al. 1984, Oviatt et al. 1986, Nixon 1992, Nixon et al. 1996. DIN was used to enrich treatment conditions (e.g. 1X...32X) and is

represented in Figures 5, 6, and 7. TN values include input of organic forms with feedwater, which is only a substantial portion of input at the

control and the low end of the enrichment gradient. Production for year 1 of experiment was extrapolated using empirical model of Keller 1988,

which did not include measurements of primary production above 600 g C m-2y

-1 (Nixon 1992). These values are used in Figures 6 and 7. 

Parenthetical production values for year 2 are from Keller 1988. Hypoxic and anoxic events were periodic, not chronic.

dExcept for Providence River, Boston Harbor and Gulf of Mexico, loading is TN as reported by Nixon et al. 1996. With noted exceptions for

individual systems below, see Nixon (1992, 1997) for productivity references.

eAlso see Elmgren 1989, Cederwall and Elmgren 1990, Rosenberg et al. 1990. Table value for TN loading from Nixon et al. 1996 is lower than

DIN input in Nixon 1997 plot, which included N input across the halocline. Lower value is labeled in 

Figure 6.

fAlso see Boynton et al. 1995, Boynton and Kemp 2000; historical Chesapeake production range (parenthetical) is from Boynton et al. 1982.
Table 2-3. DO, nutrient loading, and other characteristics for selected coastal areas and a MERL mesocosm

enrichment experiment (continued)

Nutrient Criteria—Estuarine and Coastal Waters 2-17

gMainstem stratification, increasing anoxic extent; Officer et al. 1984, Boynton and Kemp 2000.

hTop line is for dry flow, bottom line is for wet flow.

iOnly strongly stratified by freshwater at head of Bay in Providence River area, see notes j, k below. Production range is from Nixon 1997 (does

not include historical presettlement estimate of 120-130 g C m-2y

-1).

jOviatt et al. 1984, Doering et al. 1990, Asselin and Spaulding 1993; TN loading from seaward and landward inputs, avg residence time (2.5 d),

low DO in 13-15 m channel.

kUses longer 7-d residence time during very low flow conditions, Asselin and Spauling 1993.

lTN budget inclues direct estimate of ocean loading as well as land loading. Nixon et al. 1996 gave a preliminary budget; table shows improved

budget of Kelly 1998. Freshwater stratification and near hypoxia/occasional hypoxia only occur in inner harbor. See Signell and Butman 1992

for flushing estimate of whole harbor.

mNorthern harbor section, Kelly 1998. Harbor station production of Kelly and Doering 1997.

nArea represents greatest measured extent of hypoxic zone. Higher production is for immediate plume (Rabalais et al. 2000). TN loading is to a

20,000-km2 hypoxic zone only (and thus is a maximal rate) based on Mississippi/Atchafalaya input of 130 x 109 moles y-1 (Howarth et al. 1996;

Turner and Rabalais 1991). Rate is consistent with long-term average (1980-1996) estimated by CENR 2000 of 1,567,900 metric tons y-1.

oAssumed a 6-mo residence time (~seasonal turnover) for illustration only; if longer, then normalized concentration would increase accordingly.
2-18 Nutrient Criteria—Estuarine and Coastal Waters

sedimentation, horizontal transport, production, and grazing. Each forcing has its characteristic timescale

of variability (e.g., 12.4-hr tidal period, the diel 24-hr light cycle, several days to weeks-long storm

events of enhanced river flow and wind stress, and seasonal cycles of irradiance and temperature; Cloern

1996). 

Phytoplankton growth depends on nutrient supplies, as expected, but growth is significantly modulated

by complex physical processes that operate at virtually every physical scale (Giller et al. 1994). For this

reason, it is desirable for RTAGs and State water quality managers to have ready access to individuals

with a specialty in physical oceanography.

In estuaries, bottom topography and bathymetry form the basin in which tidal currents, freshwater inflow,

and wind vectors act as principal drivers of estuarine and coastal physical processes and contribute to

variability in mixing and circulation of waters (Cloern 1996) (Figure 2-8). Physical processes can

attenuate or exacerbate nutrient enrichment effects depending on the form of interaction. For example,

the Delaware River Estuary receives TN and TP loads somewhat larger than does the mainstem

Chesapeake Bay, yet the Delaware Estuary has lower phytoplankton production and does not have a

hypoxia problem, largely because of its relatively strong vertical mixing (i.e., a weak vertical density

stratification) and horizontal water exchange with the open ocean system (Pennock 1985).

Freshwater inflow is the “master driver” that defines the ecological character of river-dominated

estuaries. Boynton and Kemp (2000) proposed a simple conceptual model to explain effects of river flow

on Chesapeake Bay ecological processes associated with nutrient inputs (Figure 2-9). These authors

stated:

The importance of freshwater inputs is obvious; it is a central feature in the definition of estuarine

systems, it influences physical dynamics (Boicourt 1992), is well correlated with nutrient inputs

(Summers 1993), and has been implicated in regulating either directly or indirectly estuarine

processes ranging from primary production (Boynton et al. 1982; Cloern et al. 1983) to benthic

secondary production (Flint 1985) to fish recruitment (Stevens 1977) and catch (Sutcliffe 1973;

Sutcliffe et al. 1977; Ennis 1986).

Boynton and Kemp applied regression techniques to datasets from mid-Chesapeake Bay, a mesohaline

area, to test the ideas represented in Figure 2-9. They showed that Susquehanna River flow was

significantly related to annual average primary production, annual average surface chlorophyll a, spring

deposition of total chlorophyll a per square meter, and total chlorophyll a deposition rate (meter squared

per day). They also showed that the decline in dissolved oxygen concentrations in deep water during the

spring bloom period was also related to flow (Figure 2-10). Although this relationship could be driven

by riverflow effects on stratification, which in turn regulates dissolved oxygen depletion, they argue that

river inputs of nutrients are of primary concern. This is because years of high and low stratification did

not correlate well to years of high and low rates of oxygen decline. The implication is that nutrient

enrichment played a key role in deep-water hypoxia.
Nutrient Criteria—Estuarine and Coastal Waters 2-19

Figure 2-8. Cartoon diagrams of three physical forcings that operate at the

interface between SCEs and the coastal ocean (tides), watershed (river inflow), and

atmosphere (wind). Each physical forcing influences the growth rate of the

resident phytoplankton population through, for example, its influence on the

distribution of suspended sediments and turbidity. Each forcing also influences the

rate of vertical mixing, with riverine inputs of freshwater as a source of buoyancy

to stratify the water column and the tide and wind as sources of kinetic energy to

mix the water column. Each forcing is also a mechanism of water circulation that

transports phytoplankton horizontally. Much of the variability of phytoplankton

biomass during blooms can be understood as responses to fluctuations in these

interfacial forcings. Source: Cloern 1996.
2-20 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-9. Simple schematic diagram showing the influences of river flow on ecosystem

stocks and processes examined in this study. The mechanistic relationships between river

flow and the stocks and processes shown in the diagram are explained in the text. Source: 

Boynton and Kemp 2000.

Figure 2-10. Scatter diagram showing the relationship between the rate of

decline in dissolved-oxygen concentrations in deep water (dDO dt!1) and

average deposition rates of total chlorophyll a during the spring-bloom period. 

Data are from the 1985-1992 period and were collected at the R-64 site. The

date on which hypoxia (DO concentration <1 mg 1-1) was first encountered

during highest (1987) and lowest (1992) deposition years is also indicated.
Nutrient Criteria—Estuarine and Coastal Waters 2-21

Freshwater inflow plays a major role in the degree of stratification (Figure 2-11a-d) and nontidal flushing

(Figure 2-12) of estuaries. Density stratification influences the depth of vertical mixing relative to the

euphotic zone depth and the tendency toward hypoxia formation, that is, the effect of sealing off bottom

waters from reaeration. On a seasonal basis, stratification greatly influences the degree of hypoxia, but

seems to have a lesser role on an interannual scale (see above paragraph). Tidal displacement also

contributes to flushing (Figure 2-13). Numerous studies have documented the role of freshwater inflow

regulation of primary production through interaction with other estuarine processes via different

mechanisms (Pennock and Sharp 1994, Harding and Perry 1997, Cloern 1996, Sin et al. 1999). Freshets

deliver substantial quantities of nutrients to an estuary and lead to blooms (Mallin et al. 1993, Rudek et

al. 1991). Effects of rainfall operating on hydrographic processes have been shown to influence trophic

organization (Livingston 1997). A significant effect of episodic freshwater inflow is determining the

appropriate averaging period for reference conditions applicable to nutrient criteria development. The

issue applies to decadal wet and dry cycles as well. Water quality managers should anticipate that even

in estuaries relatively free of anthropogenic nutrient enrichment, some level of hypoxia may occur during

wet weather cycles. This “natural” condition, should it be observed will need to be factored into nutrient

criteria development.

Other Physical Factors

Other physical factors (e.g., salinity, temperature, and light) influence the expression of nutrient

enrichment effects and are extensively reported in standard textbooks. For example, salinity can

influence enrichment effects and can also influence biotic distributions (e.g., grazing populations),

primarily through the osmotic capabilities of resident organisms (Kinne 1964). Temperature and light

availability to photosynthetic organisms is obviously important. Temperature regulates, within certain

limits, the metabolic rates of organisms, especially poilkilotherms, and influences the distribution of

many species. Light also influences the feeding behavior of many planktonic animal forms, especially

crustacean filter feeders, which has relevance to algal grazing. Climatic factors influence phytoplankton

biomass production in estuaries (Lehman 2000). Additional information on the roles of temperature and

light as limiting factors to net primary production and effects of nutrient overenrichment is provided in

Appendix B.

2.3 NUTRIENT LOADS AND CONCENTRATIONS: INTERPRETATION 

OF EFFECTS

The issue of whether or not to focus on nutrient concentration versus loading criteria has been a

contentious one among both scientists and managers. Whether or not to use concentrations or

loading as criteria largely depends on the spatial and temporal scales of assessing ecosystem

responses to nutrient inputs (H. Paerl, personal communication).
2-22 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-11a–d. Schematic diagram of coastal plain estuary types, indicating direction

and degree of mixing. Arrows show direction of net mass transports of water, and the

arrow size indicates the relative magnitudes of the transports. Source: Lippson et al.

1979.
Nutrient Criteria—Estuarine and Coastal Waters 2-23

Figure 2-12. Net transports in estuaries resulting from estuarine flows and mixing. At

any one point along an estuary, the difference between upstream- and downstreamdirected transports is equal to the freshwater input to that point. In this example with no

tributaries, the difference is equal to the input at the head of the estuary. Source: Lippson

et al. 1979.

Figure 2-13. Net movement of a particle in each layer of a two-layered flow system. 

Source: Lippson et al. 1979.
2-24 Nutrient Criteria—Estuarine and Coastal Waters

Conceptual Framework

Nutrient concentrations are what phytoplankton (and other plants) respond to instantaneously or on very

short time scales. The dissolved inorganic and, to some extent, organic nutrient concentrations that

remain in a water parcel after a short period of phytoplankton growth are largely what is left over or

unused. (Note: Some dinoflagellates can obtain nutrients from particulate materials and exhibit other

complex forms of nutrition.) Nutrient uptake, including any luxuriant uptake, will be mostly converted

into organic form, given a suitable short period for growth. Thus, total concentration is a measure of the

nutrient in living form as well as any unused organic and inorganic forms. If concentrations of nutrients

are to be used as criteria, the total concentration is most likely to reflect the short-term phytoplankton

growth potential (Boynton and Kemp 2000). 

Recycling is an important aspect of phytoplankton biomass production. If nutrients in a water parcel are

all converted into algal biomass, then maintaining the algal biomass requires rapid recycling or additional

supplies to the water parcel. With loss of phytoplankton from the water column through sedimentation,

grazing and conversion of phytoplankton to animal biomass, dispersion, and advection, maintenance and

any further net primary production require new supplies of nutrients. These processes all involve longer

time scales that include seasonal and interannual considerations of ecosystem water quality (i.e., use

impairments) and habitat response. 

Examples

Some examples of regression relationships between nutrient load and concentration and response

variables are instructive because nutrient concentration often does not provide a useful relationship. 

There is a range in the lag time between nutrient load and coastal water ecosystem responses. Such lags

have been reported for a number of estuaries, including the Patuxent (Kemp and Boynton 1984),

mainstem of the Chesapeake Bay (Malone et al. 1988), mesohaline York River estuary (Sin et al. 1999),

and Logan River and Moreton Bay, Australia (O’Donohue and Dennison 1997). Nixon et al. (1996)

developed a number of regressions between residence time and response variables (e.g., percent total N,

percent P exported, percent N retained from land and atmosphere, and percent N denitrified) from a

number of estuaries and coastal marine systems. Dettmann (in press) developed relationships somewhat

similar to those of Nixon et al. that included some different estuaries and coastal waters employing a

modified algebraic expression for residence time (e.g., Figure 2-14). The temporal scale of these

regressions typically ranges from months to annual averages. These regressions help frame causal

relationships but usually are not adequate by themselves to establish nutrient criteria. For example, the

Delaware Bay lies between the northern Adriatic Sea and Chesapeake Bay in terms of the fraction of N

exported, but the Delaware Bay has few symptoms of nutrient overenrichment.

For a number of coastal embayments in Virginia and Maryland, chlorophyll a concentration regressed on

a TN loading rate that was scaled to a unit area loading rate of the receiving waterbody surface area,

resulting in a relatively high R2 (Boynton et al. 1996). Peak chlorophyll a concentrations in the Potomac

Estuary regressed against peak TN load showed the highest chlorophyll a concentrations occurred under

average flow conditions (Boynton 1997). Maximum freshwater inflows resulted in a very strong density

stratification, but the nutrients were advected into the lower Chesapeake Bay, and thus no bloom formed
Nutrient Criteria—Estuarine and Coastal Waters 2-25

Figure 2-14. The fraction of landside nitrogen input exported from 11 North

American and European estuaries versus freshwater residence time (linear time

scale). Baltic Sea not shown. Source: Dettmann (in press).

in the lower Potomac estuary. Low freshwater inflows resulted in much weaker vertical density

stratification and apparently a low nutrient supply that limited phytoplankton bloom potential (Figure

2-15). 

Using an interannual time scale, Harding (1994) summarized the historical (1950–1994) nutrient and

chlorophyll a trends for the mainstem of the Chesapeake Bay. Nitrogen, P, and chlorophyll a

concentrations increased considerably over the period of record. Harding and Perry (1997) applied a

statistical time series model and determined that confounding effects of freshwater inflow did not explain

the chlorophyll a increase in the lower bay. The DIN:DIP ratios suggested a greater influence of DIN as

a limiting nutrient to biomass production. Variation in the flow of the Susquehanna River over the

period of record tends to cloud the empirical relationships, especially in the oligohaline region and

brackish zone. 

By inference, nutrients were hypothesized to be the principal causative agent. Since the 1970s, the

winter-spring freshet has been associated with a strong diatom bloom, and in 1989 a drought delayed

delivery of DIN and Si to the mesohaline reach of the bay until late spring, thus leading to a late-season

phytoplankton biomass increase composed primarily of flagellates.

Phytoplankton growth and biomass accumulation appear to be directly related to riverborne nutrient

inputs in the Chesapeake Bay (Boynton et al. 1982, Malone et al. 1988). Typically, years with higher

river flow (within limits) are marked by greater algal biomass, which supports elevated respiration and

more rapid depletion of bottom water DO in deep, stratified estuaries (Boicourt 1992). However, this

relationship is confounded by interannual variations in salinity stratification because stratification is

directly related to river flow (Seliger and Boggs 1988, Officer et al. 1984). Distinguishing between the
2-26 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-15. Scatter plots of water column averaged chlorophyll a at a mesohaline station (MLE 2.2) versus

several different functions of total nitrogen (TN) loading rate measured at the fall line of the Potomac River

estuary. Source: Boynton 1997.
Nutrient Criteria—Estuarine and Coastal Waters 2-27

effects of physical and biological processes on interannual variations in anoxia/hypoxia is only now

beginning on the basis of mathematical modeling and long-term empirical monitoring data. Stratification

from freshwater inflow from the Susquehanna River apparently is insufficient by itself to explain the

increased hypoxic volumes in the Chesapeake Bay from the early 1980s to 1999 (J. Hagy, personal

communication). In shallow estuaries the hypoxic volume, if present, is likely to be highly variable

spatially owing to the influence of variable freshwater inputs and estuarine in situ physical factors that

cause wide excursions and mixing of water masses (e.g., Neuse River estuary, H. Paerl, personal

communication).

A detailed study of nutrient and phytoplankton relationships in the mesohaline region of the mainstem of

the Chesapeake Bay demonstrated that “despite high inputs of DIN and dissolved silicate relative to DIP

(molar ratios of N:P and Si:P > 100), seasonal accumulations of phytoplankton biomass within the saltintruded reach of the bay appear to be limited by DIN supply while the magnitude of the spring diatom

bloom is governed by the dissolved Si supply” (Malone et al. 1996, Conley and Malone 1992). The

maximum chlorophyll-specific productivity occurred in the late summer, the maximum biomass occurred

in the spring, and volumetric-based productivity occurred in midsummer (see their Figure 4). This

temporal asymmetry leads to difficulties in ascribing simple empirical relationships between

phytoplankton biomass and nutrient concentrations.

2.4 PHYSICAL-CHEMICAL PROCESSES AND DISSOLVED OXYGEN DEFICIENCY

Dissolved oxygen deficiency, or hypoxia, is of critical importance to the health of aquatic life. The role

of physical processes, especially mixing and physical circulation of estuarine waters, has been widely

reported in the literature (Smith et al 1992). “There is no other environmental variable of such ecological

importance to coastal marine ecosystems that has changed so drastically in such period of time as

dissolved oxygen” (Diaz and Rosenberg (1995). One of the earliest studies to measure DO in a U.S.

estuary occurred in the Chesapeake Bay and Potomac River in 1912 (Sale and Skinner 1917),

approximately two decades after Winkler developed his now legendary method for determining the

concentration of DO in aquatic systems. Hypoxia was already present in the bottom waters of the lower

Potomac River estuary at this early date because a measurement indicated only a DO < 2.0 ml/L, or 35%

saturation. 

Individual species exhibit a range in adaptability to relatively low DO concentrations (e.g., see “EPA

822-D-99-002 Draft Ambient Water Quality Criteria for Dissolved Oxygen [Saltwater]: Cape Cod to

Cape Hatteras”). Hypoxia and H2S apparently cause synergetic effects that make marine benthic animals

more sensitive to hypoxia when H2S is present (Diaz and Rosenberg 1995). These authors suggest that

the occurrence of hypoxia in shallow coastal and estuarine areas appears to be increasing, and evidence

suggests that the increase has global dimensions and seems most likely to be accelerated by human

activities (Nixon 1995, Bricker et al. 1999). Although hypoxia has undesirable consequences, when

bottom waters go anoxic wholesale biogeochemical changes occur. These changes can include release of

phosphate from sediments, emergence of highly toxic hydrogen sulfide, elimination of nearly all
2-28 Nutrient Criteria—Estuarine and Coastal Waters

multicellular animals from sediment habitats, reduction in the coupled nitrification-denitrification, and

changes in metal solubilities, with many metals becoming toxic. 

Diaz and Rosenberg (1995) concluded that should DO concentrations become slightly lower, catastrophic

events may overcome the systems and alter the productivity base that leads to economically important

fisheries and amenities. Aquatic biota exposed to low DO concentrations may be more susceptible to the

adverse effects of other stressors such as disease, toxic chemicals, and habitat modification (Holland

1977). Low DO conditions can increase the vulnerability of the benthos to predation, as the infaunal

animals extend above the sediment surface to obtain more oxygen (Holland et al. 1987). Dissolved

organic carbon apparently is a major carbon and energy source for bacteria (i.e., microbial loop; Azam et

al. 1983), whose metabolism is a major cause of hypoxia. Hypoxia and anoxia indicate that a coastal

ecosystem is severely stressed by nutrient overenrichment and should receive immediate attention by

water quality managers.

2.5 NUTRIENT OVERENRICHMENT EFFECTS AND IMPORTANT 

BIOLOGICAL RESOURCES

Benthic Vascular Plant Responses to Nutrients

A major lesson learned over the past 25 years is that nutrient overenrichment has had a devastating effect

on SAV, whether estuarine species or higher salinity seagrasses. This conclusion is based on work

conducted mostly on the U.S. Gulf of Mexico and Atlantic Coasts (Tomasko et al. 1996, Tomasko and

LaPointe 1991, Kemp et al. 1983, Orth and Moore 1983, Burkholder et al. 1992, Taylor et al. 1995, Short

et al. 1995). Dennison et al. (1993) reported the following habitat criteria for SAV: DIN of 10.7 μM,

DIP of 0.33 μM; N:P (atomic) of 32; and chlorophyll a of 15 μg/L. These criteria are being re-analyzed

by the EPA Chesapeake Bay Program. 

The relationship between N load and concentration and chlorophyll a is not limited to phytoplankton. 

Predictive regression relationships between N and chlorophyll a, water column light attenuation, and

seagrass recovery in Tampa Bay were found for N loading, not ambient N concentrations (Janicki and

Wade 1996, Greening et al. 1997). Tomasko et al. (1996) detected a negative correlation between N

loads and turtle grass (Thalassia testudinum) biomass and productivity in Sarasota Bay, FL. 

Moore and Wetzel (2000) determined experimentally that eelgrass (Zostera marina) in the York River

estuary, lower Chesapeake Bay, is exposed to N concentrations adequate to stimulate enough epiphytic

growth to shade out this vascular plant. In mesocosms containing a complex of species characteristic of

shallow marine coastal lagoons along the Narragansett Bay coast, Taylor et al. (1995) showed that N

alone—but not P alone—caused an increase in water column concentrations of chlorophyll a and

particulate N, increased daytime net production, and increased growth of juvenile winter flounder. 

Eelgrass beds and drift algae apparently were shaded out by phytoplankton at high nutrient levels. 

Experiments conducted by Neundorfer and Kemp (1993) on the submersed plant Potamogeton

perfoliatus in microcosms using lower Choptank Estuary water demonstrated that effects of N and P on

algal densities were synergistic in that responses to N addition were greatest at high P loading and vice
Nutrient Criteria—Estuarine and Coastal Waters 2-29

versa. Also, combined amendments (N+P) at highest treatment rates resulted in epiphytes and

phytoplankton increasing more than when these nutrients were added individually. On the basis of

microcosm studies and the literature, Sturgis and Murray (1997) suggested that there may be a more

complex relationship between nutrient enrichment and SAV growth and survival. For example, the

relationship may depend on the form, delivery frequency, and loading rate of nutrients.

There now appears to be enough scientific data and knowledge to establish nutrient regimes that will

protect temperate and subtropical seagrass ecosystems. 

Other Examples of Important Biotic Effects of Nutrient Overenrichment

It is difficult to find recent quantitative relationships between nutrient loading and fishery impacts for

coastal systems. One explanation is that the large marine vertebrate species which are mostly extinct or

severely over-fished help determine the nutrient assimilative capacity of marine ecosystems including

estuaries and coastal waters (Jackson et al. 2001). For economically important fisheries, variable fishing

pressure may cloud the analysis and other factors may vary to obscure nutrient-related patterns. Often,

one is left with mostly anecdotal insights as to potential negative effects of overenrichment on higher

trophic levels focusing on data and insights only from recent decades. There is a plausible and positive

relationship between marine fisheries yield and nitrogen supply, with a wide range in estuarine and

coastal marine habitats represented (Nixon 1992). This approximately natural response is analogous to

what mariculturists attempt to achieve when they fertilize fish enclosures, but these enclosures, whether

on land or in the marine environment, are known to cause local water quality problems. The relationship

Nixon reported on involved a two-step function: a positive relationship between primary production (g C

m-2 y-1) and DIN input (moles m-2 y-1) and between fisheries yield (kg ha-1 y-1) and primary production

(Figure 2-16a-c). In contrast to the foregoing positive relationship, a pelagic-demersal ratio from fishery

landings from 14 study areas in European coastal waters appeared to be a proxy for the differential

impact of nutrients on pelagic and benthic systems mediated by nutrient enrichment, resulting in hypoxia

(de Leiva Moreno et al. 2000). A general model suggests that overenrichment can lead to decreased

fisheries productivity (Figure 2-17).

Oysters are ecosystem engineers that create biogenic reef habitat important to estuarine biodiversity,

benthic-pelagic coupling, and fishery production (Lenihan and Peterson 1998). These authors conducted

an analysis of habitat degradation (i.e., oyster dredging) through fishery disturbance that enhanced

impacts of hypoxia on oyster (Crassostrea virginica) reefs in North Carolina. This is a fairly

complicated story but the conclusions from the analysis seem inescapable. Dredging lowered the oyster

reef into the hypoxic zone where the reef and associated organisms died from DO depletion. Another

example of effects of nutrient overenrichment causing impacts on oysters was reported by Ryther (1954)

for Long Island, New York duck farms where nutrient enrichment caused phytoplankton to grow that

were indigestible for oysters.

Hypoxia is known to kill other benthic organisms. Diaz and Rosenberg (1995) cited many studies where

hypoxia resulted in the deaths of benthic communities. A related cause with hypoxia is that polychaetes

may extend themselves out of their sediment burrows and become easier prey to fish predators. Another 
2-30 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-16b. Primary production by phytoplankton (14C uptake) as a function of the annual

input of dissolved inorganic nitrogen per unit area of a wide range of marine ecosystems. 

Source: Nixon (1992).

Figure 2-16a. Primary production by phytoplankton (14C uptake) as a function of the estimated annual

input of dissolved inorganic nitrogen per unit volume of a wide range of marine ecosystems. Source: 

Nixon (1992).
Nutrient Criteria—Estuarine and Coastal Waters 2-31

Figure 2-16c. Fisheries yield per unit area as a function of primary production in a wide range of estuarine

and marine systems. Modifed from Nixon (1988) to include a revised primary production estimate for the

Peru Upwelling from Guillen and Calienes (1981). Systems identified and data sources in Nixon (1982)

and Nixon et al. (1986). Source: Nixon (1992).
2-32 Nutrient Criteria—Estuarine and Coastal Waters

Figure 2-17. Comparative evaluation of fishery response to nutrients. Although higher

nutrient concentrations initially increase the productivity of fisheries, ecological systems

worldwide show negative effects as nutrient loading increases and hypoxic or anoxic

conditions develop. Each generic curve in the lower half of the figure represents the

reaction of a species guild to increasing nutrient supplies. The top half of the figure

illustrates trends in various marine systems around the world. Reversals show that trends

toward overenrichment have been turned around in several areas. Source: CENR 2000.

effect of hypoxia on the biota is the loss of sufficient bottom habitat. This is often difficult to

quantitatively relate to economically important species but the negative effect may still be real. If

endangered species are present, this hypoxic effect is one of direct societal and legal concern.

2.6 CONCLUDING STATEMENT ON NITROGEN AND PHOSPHORUS CONTROLS

It is important to note that in estuaries and nearshore coastal marine waters, the fact that nitrogen often

limits algal biomass production does not mean that managers should be unconcerned about phosphorus

enrichment. In river-dominated temperate estuaries, the upper reaches of estuaries, such as lakes and

rivers, are often phosphorus limited. The manager who therefore concentrates on phosphorus

management alone risks letting an undue amount of nitrogen proceed downstream to exacerbate problems 
Nutrient Criteria—Estuarine and Coastal Waters 2-33

where an abundance of P allows the excess N to drive trophic conditions to unacceptable levels of

nutrient enrichment.

Similarly, any reductions achieved in P loadings and concentrations at the coastal margin will limit

potential eutrophy/hypertrophy even in the face of abundant nitrogen. Consequently, the prudent

management strategy is to limit both phosphorus and nitrogen. Emphasis on one or the other as an

element of symptomatic management in fresh or saline waters may be appropriate in some cases, but the

manager must always be concerned about the downstream consequences and the net enrichment effects

to the larger system. 

In summary, attempting to understand the nutrient overenrichment problem in estuaries and coastal

ecosystems primarily from a bottom-up perspective provides a limited perspective. This manual has

included references to the historical past that reported on potential positive effects of top-down controls

on nutrient overenrichment. It is likely that the most scientifically robust nutrient criteria will need to

take into account the effects of past overfishing and its consequences for marine eutrophication (Jackson

et al. 2001). Thus, higher trophic levels are more than just a thermodynamic response to nutrient

enrichment because they help modulate many of the negative consequences of overenrichment.

Ecological feedback mechanisms that involve higher trophic levels can be a positive tool in nutrient

management.
Nutrient Criteria—Estuarine and Coastal Waters 3-1

Major Factors Influencing Estuarine Susceptibility

 to Nutrient Overenrichment

Examples of Coastal Classification

Coastal Waters Seaward of Estuaries

CHAPTER 3

Classification of Estuarine and

Coastal Waters

3.1 INTRODUCTION

Purpose and Background

Classification is an important step in addressing the problem of degradation, especially because of

nutrient overenrichment. There are too many nutrient-degraded estuaries in the United States for the

Nation to conduct comprehensive ecosystem studies of all those affected by overenrichment. Where

possible, similar estuaries, tributaries, or coastal reaches should be equated through physical

classification to reduce the magnitude of the criteria development problem and to enhance predictability

of management responses. To be useful, classification should reduce variability of ecosystem-related

measures (e.g., water quality factors) within identified classes and maximize interclass variability. This

is important because managers need to understand how different types of estuaries and coastal waters, as

well as important habitat differences within these systems, respond to nutrient overenrichment in order to

plan effective management strategies. 

The ecosystem processes that regulate nutrient dynamics, discussed in Chapter 2, should provide the

elements for initial development of a useful classification system. Although predicting susceptibility of

estuarine and coastal waters to nutrient overenrichment is in a primitive state, several approaches are

reviewed because they have some utility even if they are only marginally adequate for prediction of

nutrient effects. The general approach is also appropriate for coastal systems.

General trends relate N loading with chlorophyll and primary productivity; however, these trends are

seldom usefully predictive for individual systems or for all classes of coastal systems (Kelly in press). 

Progress has been especially slow in predicting many of the secondary, but societally important, effects

of nutrient overenrichment, e.g., bottom water dissolved oxygen (DO) deficiency, harmful algal blooms

(HABs) or species-specific HABs, formation of macroalgal mats, fisheries productivity, and species

composition. For many cross-system comparisons, N loading and SAV decline have become more

predictive than for other indirect effects (Duarte 1995; Dennison et al. 1993), but even here the

predictions may be confounded by highly variable ecosystem factors. A major impedance to effective

understanding is limited comparative studies designed to test hypotheses regarding estuarine

susceptibility to nutrient enrichment (Turner 2001). Post hoc comparative approaches and assessment of

disparate studies have been useful but clearly inadequate (Livingston 2001b). 

Post hoc statistical approaches have helped explain some of the variability in eutrophication, but have not

captured the actual mechanisms and their interactions controlling eutrophication across estuaries and
3-2 Nutrient Criteria—Estuarine and Coastal Waters

coastal waters (NRC 2000). The ability to explain the mechanisms in a predictive manner is clearly a

critical national and global research need, as nutrient overenrichment of coastal ecosystems extends far

beyond the shores of the United States. For site-specific criteria, several approaches are available,

including empirical regression and mechanistic simulation models. The large effort typically required to

calibrate and verify mechanistic models is an indicator of the difficulty in understanding the many

potential confounding factors of ecosystem-level prediction. A basic premise of this manual is that

knowledge of the physical setting and the minimally disturbed ecosystem reference condition must

underpin monitoring and management efforts to protect and restore coastal systems impaired by nutrient

overenrichment. 

Only in approximately the past two decades have comparative studies of nutrient dynamics among two or

more relatively large estuaries been published (e.g., Fisher et al. 1988; Malone et al. 1999; Pennock et al.

1994). Comparative analysis of the Delaware and Chesapeake Bays has provided insights regarding

processes that control expression of nutrient enrichment (Chapter 2). For example, both systems are

drowned river mouth coastal plain estuaries and are located adjacent to each other along the coast, but

have very different responses to nutrient loading. Delaware Bay has a somewhat larger nutrient load than

the Chesapeake, but has few of the nutrient enrichment symptoms well chronicled for the Chesapeake

Bay (Flemer et al. 1983, Sharp et al. 1994, Chesapeake Bay Program Periodic Status Reports). Similar

insights have been provided by comparing nutrient processes between Delaware and Mobile Bays

(Pennock et al. 1994). Susceptibility appears to be largely explained by differences in the physics of

flushing, including bathymetry and related physical habitat differences.

Defining the Resource of Concern

As a first step in classification, defining the resource of concern is important. Resources of concern are

estuaries and coastal waters located in the contiguous States or within authorized Tribal lands. Managers

must decide which waterbodies to include in the population to which criteria will be applicable. A lake

classification may exempt small ponds that might be excluded because of their size and man-made

nature, whereas tidal creeks, although small, still have a functional connection to the larger estuary and

might not be excluded because of size. Many estuaries and coastal waters share multiple political

boundaries, and for the sake of consistency all involved jurisdictions should jointly decide on the scale of

inclusiveness. For open coastal waters, a State or authorized Tribe’s legal authority may extend for a

relatively short distance on the continental shelf, e.g., 3 nautical miles. However, coastal oceanographic

processes seaward of the statutory limit likely influence nutrient overenrichment processes and

exacerbate the difficulty of diagnosing the anthropogenic contribution to nutrient problems.

3.2 MAJOR FACTORS INFLUENCING ESTUARINE SUSCEPTIBILITY TO

NUTRIENT OVERENRICHMENT

The NRC (2000) publication summarized approximately a dozen factors deemed important to

characterize the susceptibility of estuaries to nutrient loading. A short list is provided; however, it is

expected that the following list will be modified and refined as more is learned about the subject:
Nutrient Criteria—Estuarine and Coastal Waters 3-3

1. System dilution and water residence time or flushing rate

2. Ratio of nutrient load per unit area of estuary

3. Vertical mixing and stratification

4. Algal biomass (e.g., chlorophyll a, and chlorophyll a corrected for nonchlorophyll a light

attenuation over seagrass/SAV beds and macroalgal biomass as AFDW)

5. Wave exposure (especially relevant to seagrass potential habitat)

6. Depth distribution (bathymetry and hypsographic profiles)

7. Ratio of side embayment (s) volume to open estuary volume or other measures of embayment

influence on flushing. 

Several terms listed above are briefly discussed because their significance often is not adequately

appreciated. 

Dilution

The volume of an estuary affects its ability to dilute inflowing nutrients. Thus, the loading rate of

nutrient per unit volume of the estuary is a better indicator of the potential for exceeding the assimilative

capacity of the estuary as a whole than is the absolute loading rate. This ratio may not express the

potential for local effects near the point of entry into the estuary, as nutrients there are diluted by only a

fraction of the total estuary volume. The potential for such local effects is reduced if mixing into the

main body of the estuary is rapid. 

Water Residence Time

Estuaries that flush rapidly (i.e., have a short residence time) will export nutrients more rapidly than

those that flush more slowly, resulting in lower nutrient concentrations in the estuary. Dettmann (in

press) has derived a theoretical relationship between the mean residence time of freshwater in an estuary

and the increase in the average annual concentration of total nitrogen in the estuary as a result of inputs

from the watershed and atmosphere. In addition, estuaries with residence times shorter than the doubling

time of algal cells will inhibit formation of algal blooms. Residence time or flushing rate is discussed in

more detail in Appendix C.

Stratification

Highly stratified systems are more prone to hypoxia than are vertically mixed systems. Stratification not

only limits downward transport of oxygen from atmospheric reaeration, it also retains nutrients in the

photic zone, making them more available to phytoplankton. In stratified systems, it may be more

appropriate to estimate the dilution potential of the estuary using the volume above the pycnocline rather

than the entire volume of the estuary.

It is expected that the shortened list will be revised and modified as more is learned about factors

important in estuarine and coastal waters classification. Some of these factors will apply to estuaries and

others to coastal waters. 
3-4 Nutrient Criteria—Estuarine and Coastal Waters

3.3 EXAMPLES OF COASTAL CLASSIFICATION 

Scientists and resource managers have used various classification schemes for many years to organize

information about ecological systems. As discussed earlier, estuaries and coastal water systems are

characterized by a suite of factors (e.g., river flow, tidal range, basin morphology, circulation, and

biological productivity) that are ultimately controlled largely by geology and climate. A review of

Chapter 6 in the NRC (2000) publication provides useful descriptions of the various approaches to

estuarine classification, and some pertinent features are highlighted below. 

Geomorphic Classification

Geomorphic classification schemes provide some insight into the circulation structure and are a firstorder estimate of water residence time or flushing characteristics. Such classifications may not in

themselves be predictive of susceptibility to nutrient enrichment, but they are a useful place to begin a

first-order assessment of susceptibility. Knowledge of deep channels, however, identifies potential areas

subject to hypoxia, and the extent of shallow waters and associated factors (e.g., wind fetch) often

provides insights into potential seagrass habitat.

Estuaries can be divided geomorphically into four main groups (Pritchard 1955, 1967; Dyer 1973): (1)

coastal plain estuaries, (2) lagoonal or bar-built estuaries, (3) fjords, and (4) tectonically caused estuaries. 

This classification frequently appears in textbooks, and only some important features relative to nutrient

susceptibility are described.

Coastal Plain Estuaries: Classical and Salt Marsh

Both subclasses are characterized by well-developed longitudinal salinity gradients that influence

development of biological communities. Examples of the classical type include the Chesapeake Bay (the

largest estuary of this type), Delaware Bay, and Charleston Harbor, SC. Vertically stratified systems

with relatively long residence times (e.g., Chesapeake Bay) tend to be susceptible to hypoxia formation. 

Pritchard (1955) further classified drowned river valley estuaries into four types (A-D) depending on the

advection-diffusion equation for salt (Table 3-1). Type C estuaries are less sensitive to algal bloom

formation and hypoxia because of mixing features.

The salt marsh estuary lacks a major river source and is characterized by a well-defined tidal drainage

network, dendritically intersecting the extensive coastal salt marshes (Day et al. 1989). Exchange with

the ocean occurs through narrow tidal inlets, which are subject to closure and migration following major

storms (e.g., Outer Banks, NC). Consequently, salt marsh estuarine circulation is dominated by

freshwater inflow, especially groundwater, and tides. The drainage channels, which seldom exceed a

depth of 10 m, usually constitute less than 20% of the estuary, with the majority consisting of subaerial

and intertidal salt marsh. These systems are a common feature of the Atlantic coast, particularly between

Cape Fear, NC, and Cape Canaveral, FL. Mangrove estuaries occur from around Cape Canaveral south

on Florida’s east coast and on Florida’s west coast from around Tarpon Springs south. Nutrient

dynamics, primary production, and system respiration that occur within emergent marshes may greatly

affect water quality in the estuarine channels (Cai et al. 1999).
Nutrient Criteria—Estuarine and Coastal Waters 3-5

Table 3-1. General drowned river valley estuarine characteristics

Estuarine 

typea

Dominant 

mixing force

Mixingenergy 

Width/depth

ratio

Salinity

gradient

Mixingindexb

Turbidity Bottom

stability

Biologicalproductivity

Example

A River flow Low Low Longitudinal 

vertical

1 V. high Poor Low Southwes

Mississipp

B River flow, tide Moderate Moderate Longitudinal 

vertical

<1/10 Moderate Good V. high Chesapea

C Tide, wind High High Longitudinal 

lateral

<1/20 High Fair High Delaware 

D Tide, wind V. high V. High Longitudinal ? High Poor Moderate ?

aFollows Pritchard’s advection-diffusion classification scheme.bFollows Schubel’s definition: MI = equation here (vol. freshwater discharge on 1⁄2 tidal period) / (vol. tidal prism).

Source: Neilson and Cronin, 1981.
3-6 Nutrient Criteria—Estuarine and Coastal Waters

Lagoons

Lagoons are characterized by narrow tidal inlets and are uniformly shallow (i.e., less than 2 m deep)

open-water areas. The shallow nature enhances sediment–water nutrient cycling. Flushing is typically 

of long duration. Most lagoonal estuaries are primarily wind-dominated and have a subaqueous drainage

channel network that is not as well drained as the salt marsh estuary. Lagoons fringe the coast of the

Gulf of Mexico and include the mid-Atlantic back bays; Pamlico Sound, NC; and Indian River Lagoon,

FL. Although these systems are typically shallow, they may have pockets of hypoxic water subject to

spatial variability because of freshwater pulsing and wind effects. Some lagoonal systems have relatively

strong vertical stratifications near the freshwater river mouth and may be subject to hypoxia formation

(e.g., Perdido Bay, AL/FL; Livingston 2001a).

Fjords and Fjordlike Estuaries

Classical fjords typically are several hundred meters deep and have a sill at their mouth that greatly

impedes flushing. Hypoxia/anoxia is often a natural feature but anthropogenic nutrient loading can

severely exacerbate the problem. Examples of classical fjords on the North American continent can be

found in Alaska and Washington State (Puget Sound). Some other estuaries were also formed by glacial

scouring of the coast, but in regions with less spectacular continental relief and more extensive

continental shelves. Examples of these much shallower, fjordlike estuaries can be found along the Maine

coast.

Tectonically Caused Estuaries

Tectonically caused estuaries were created by faulting, graben formation (i.e., bottom block-faults

downward), landslide, or volcanic eruption. They are highly variable and may resemble coastal plain

estuaries, lagoons, or fjords. San Francisco Bay is the most studied estuary of this type (Cloern 1996).

Man-Made Estuaries

Especially around the Gulf of Mexico, dredged bayous, canals, and salt water impoundments with weirs

function as estuaries but do not fit well any of the other types presented. As a special case, especially in

the Gulf of Mexico, the passes of some estuaries periodically were closed off by storms and historically

remained closed until a natural event reopened them (e.g., Perdido Bay, AL/FL; R. Livingston, personal

communication). In recent years, these systems typically are maintained in an open condition by

dredging. Dredged inlets such as at Ocean City, MD, also fit this classification.

Physical/Hydrodynamic Factor–Based Classifications

Classification Using Stratification, Mixing, and Circulation Parameters

Estuarine circulation was a dominant consideration used in an earlier classification of the Chesapeake

Bay, a coastal plain system, and major tributaries, and is largely utilized today with some modifications

(Flemer et al. 1983). Coastal plain estuaries are sometimes classified by mixing type: highly stratified,

partially mixed (moderately stratified), or well mixed (vertically homogeneous). The flow ratio of these

estuaries (the ratio of the volume of freshwater entering the estuary during a tidal cycle to its tidal prism)

is a useful index of the mixing type. If this ratio is approximately 1.0 or greater, the estuary is normally
Nutrient Criteria—Estuarine and Coastal Waters 3-7

highly stratified; for values near 0.25 the estuary is normally partially mixed; and for ratios substantially

less than 0.1, it is normally well mixed (Biggs and Cronin 1981).

Stratification/Circulation Parameters

Hansen and Rattray (1966) developed a two-parameter classification scheme based on circulation and

stratification of estuaries. Circulation is described by the nondimensional parameter Us/Uf

, where Us is

the net (time-averaged) longitudinal surface current and Uf

 is the cross-sectional average longitudinal

velocity. Stratification is represented by the nondimensional parameter 
S/So, where 
S is the top-tobottom difference in salinity and So is the mean salinity. Jay et al. (2000) review alternative twoparameter classification schemes involving parameters such as the ratio of tidal amplitude to mean depth,

along-estuary and vertical density differences and vertical tidal excursion of isopycnals, or other factors

that take into account effects of tidal flats and provide additional discussion to which the reader is

referred for additional insights. They argue that the merit of the approach is its simplicity of parameters

employed and the predictive ability with regard to salt transport needed to maintain salt balance in

modeling.

Classification Using Water Residence Time

Water residence time, the average length of time that a parcel of water remains in an estuary, influences a

wide range of biological responses to nutrient loading. The residence time of water directly affects the

residence time of nutrients in estuaries, and therefore the nutrient concentration for a given loading rate,

the amount of nutrient that is lost to internal processes (e.g., burial in sediments and denitrification), and

the amount exported to downstream receiving waters (Dettmann in press, Nixon et al. 1996). Residence

times shorter than the doubling time of algae will inhibit bloom formation because algal blooms are

exported from the system before growing to significant numbers. Residence time can also influence the

degree of recruitment of species reproducing within the estuary (Jay et al. 2000).

There are a number of definitions of water residence time, including freshwater residence time and

estuarine residence time (Hagy et al. 2000; Miller and McPherson 1991), each with its own interpretation

and utility. Freshwater residence time is the mean amount of time required for freshwater entering the

estuary to exit the seaward boundary, whereas estuarine residence time is the average residence time in

the estuary for all water, regardless of its origins. Because nutrient loading is generally associated with

freshwater inputs, freshwater residence time is generally the most useful measure in considering estuary

sensitivity to nutrient loading. Freshwater residence time of a given estuary is influenced by numerous

factors, including freshwater loading rate (Pilson 1985; Asselin and Spaulding 1993; Hagy et al. 2000),

tidal range, and wind forcing (Geyer 1997), and therefore varies over a range of time scales. 

Residence time and volume together may be used to scale nitrogen loading to estuaries to permit

calculation of nitrogen concentrations and perform cross-system comparisons. Dettmann (in press) uses

a model that includes mechanistic representations of nitrogen export and loss within estuaries to show

that [Nu ], the contribution to the annual average concentration of total nitrogen in an estuary from upland

sources (watershed, direct discharges, and atmosphere), may be calculated as
3-8 Nutrient Criteria—Estuarine and Coastal Waters

where Ll is the annual average loading rate (mass/month) of total nitrogen from all upland sources

(watershed and atmosphere), Jfw is the freshwater residence time in months, V is the estuary volume, and

" is a parameter (value = 0.3 month-1) related to losses of nitrogen to processes such as denitrification and

burial in sediments within the estuary. 

Definitions of residence time, the methods used to measure or calculate them, variability of residence

time, and other estimators of residence time are described further in Appendix C.

River Flow, Tides, and Waves

Dronkers (1988) proposed an estuarine classification that distinguished various types of estuarine

ecosystems based on water exchange processes (e.g., river flow, tides, and waves) that greatly affect

energy and material fluxes including mixing (Table 3-2). This classification suggests that river flow in

partially mixed estuaries is essentially neutral, but its variation relative to hydrodynamic residence time

can be important in interpreting property-salinity diagrams (Cifuentes et al. 1990) (Figure 3-1). River

flow in the partially mixed mainstem of the Chesapeake Bay is seasonally important. 

Tidal Amplitude—A Dominant Physical Factor

Tidal amplitude provides a means to broadly classify estuaries relative to their sensitivity to nutrient

supplies. Monbet (1992) analyzed phytoplankton biomass in 40 estuaries and concluded that macrotidal

estuaries (mean tidal range 2 m) generally exhibit a tolerance to nitrogen pollution despite high loadings

originating from freshwater outflows (Figures 3-2a, b). These systems generally exhibit lower

concentrations of chlorophyll a than do systems with lower tidal energy, even when they have

comparable concentrations of nitrogen compounds. Estuaries with mean annual tidal ranges 2 m seem

more sensitive to dissolved nitrogen, although some overlap occurs with macrotidal estuaries.

NOAA Scheme for Determining Estuarine Susceptibility

NOAA (Bricker et al. 1999) developed a categorical approach based on surveys and decision rules that

led to a classification of estuarine nutrient export potential (e.g., dilution potential and flushing

potential). From this information a susceptibility matrix was constructed. The low, moderate, and high

susceptibility indices were combined with low, moderate, and high human levels of nutrient input,

resulting in a final matrix of overall human influence (see Appendix D for details).

Comparative Systems Empirical Modeling Approach

The empirical regression method can be used to determine the response of estuarine systems to nutrient

loading. This approach requires that the response factor be common to all systems in the analysis and

assumes that any graded response among systems is due to a common form of disturbance, e.g., nutrients. 

The space-for-time paradigm (Pickett 1988) posits that relationships between nutrient inputs and
Nutrient Criteria—Estuarine and Coastal Waters 3-9

Table 3-2. Classification of coastal systems based on relative importance of river flow, tides, and

waves to mixing 

Type River flow Tide Waves Description

I   River delta

II  River delta (plus barriers)

III  Tidal river delta

IV 0  Coastal plain estuary

V  Tidal lagoon

VI   Bay

VII   Coastal lagoon

Plus and minus designations indicate relative impacts; e.g.,  means that river discharge is very small relative to tidal

and wave energy.

Source: Adapted from Dronkers 1988.

ecologically meaningful estuarine responses, using multiple systems, have predictive capability, at least

for the systems used in the model development. This allows for a wide range in nutrient loading and

estuarine types to be included. The comparative-systems empirical approach has been used to determine,

for example, relationships between nutrient inputs and fish yields (Lee and Jones 1981, Nixon 1992),

benthic biomass, production and abundances (Josefson and Rasmussen 2000), summer ammonia flux

(Boynton et al. 1995), chlorophyll a concentration (Boynton et al. 1996, Boynton and Kemp 2000,

Monbet 1992), primary productivity (Nixon et al. 1996), and the dominant source of primary productivity

(Nixon et al. in press). In many of these cases, important environmental factors such as flushing time and

depth are used to normalize the nutrient loading in a similar way as Vollenweider (Vollenweider 1976)

Did for lakes to yield more precise relationships. Appendix E provides additional details.

Other Considerations

Habitat Type

The presence and extent of different habitat/community types may help distinguish one or more estuaries

within a region. These types may include seagrasses, mangroves, mudflats, deep channels, oyster reefs,

dominance of sand versus mud bottoms, extensive emergent marshes (typically coastal plain systems),

and the presence of unconsolidated versus rocky shorelines. Some of these categories may be

subclassified by salinity ranges (e.g., oligohaline, mesohaline, and polyhaline). Although related more to

water quality, blackwater versus turbid versus relatively clear estuaries defines a group representative of

estuaries around the Gulf of Mexico.

Theoretical Considerations

Coastal zone managers may wish to consider more theoretical approaches to classification as ecosystem

science develops a more in-depth understanding of ecosystem processes for estuaries under their

purview. Several different approaches are described in Appendix F.
3-10 Nutrient Criteria—Estuarine and Coastal Waters

Figure 3-1. Idealized micronutrient-salinity relations showing concentration and mixing

of nutrient-rich river water with nutrient-poor seawater. Source: Peterson et al. 1975. A. 

Expected concentration-salinity distribution of a substance behaving in a conservative

manner (e.g., chloride) in an estuary. B. Expected concentration-salinity distribution of a

substance for which the estuary is a source (e.g., particulate carbon). C. Expected

concentration-salinity distribution of a substance for which the estuary is a sink (e.g.,

phosphorus). D. Expected concentration-salinity distribution of a substance for which

the estuary is a pronounced sink, that is, where the concentration of the substance in the

estuary is lower than the river and the ocean (e.g., Si). Source: Biggs and Cronin 1981.
Nutrient Criteria—Estuarine and Coastal Waters 3-11

Figure 3-2a. Relationship between the mean annual loadings of dissolved inorganic nitrogen

(DIN) and the mean annual concentration of chlorophyll a in microtidal and macrotidal estuaries.

Summary

Various ways are available to classify estuaries regarding their vulnerability to nutrient enrichment. 

None appear to provide all the information a resource manager may want for decisionmaking. The

NOAA estuarine export potential (EXP) appears to have the current greatest utility for predictive

purposes for large systems, but even this approach embodies considerable variability (e.g., see Figures 6-

5 and 6-6 in NRC 2000). For embayments within a larger estuary, the comparative empirical modeling

approach has been demonstrated to have considerable utility. The more theoretical models eventually

may provide greater predictive power, especially as to biological sensitivities to nutrient enrichment. 

They are data intensive and may become more useful at a future time.

3.4 COASTAL WATERS SEAWARD OF ESTUARIES

Several approaches are available to classify coastal waters. The geomorphic focus is a good place to

begin, hydrographic considerations should follow, and finally habitat and community features should be

considered. Although functional considerations and theoretical indices are not described for coastal

waters, they have as much relevance for these waters as they do for estuaries. Even though much of the

concern for coastal waters will be within 20 nautical miles of shore, and most of that within the 3-mile

limit, elements of the following large-scale classification scheme will have value to the manager and

investigator.
3-12 Nutrient Criteria—Estuarine and Coastal Waters

Figure 3-2b. Relationship between the mean annual concentrations of dissolved inorganic nitrogen (DIN)

and chlorophyll a in microtidal and macrotidal estuaries. Source: Monbet 1992.

Geomorphic Classification

The flow of energy and nutrients through coastal food webs differs greatly among continental shelves,

and is driven largely by differences in the form and amount of primary production (e.g., seagrasses are

important in the Big Bend area of Florida and kelp forests are important habitats along much of the U.S.

Pacific Coast and sections of coastal Maine). These differences in turn ultimately are determined by

differences in local and ocean-scale patterns of climate (e.g., light and temperature effects), water

circulation, chemistry, and shelf geomorphology (Alongi 1998). The spring bloom, especially along the

U.S. Atlantic Coast, generally progresses from low to higher latitudes but with sharper seasonal peaks

toward higher latitudes. Variability in the progression should be considered in any classification scheme. 

Because near-coastal shelf oceanographic processes usually are not limited by the jurisdiction of a single

State, it is important that a similar classification approach be shared among coastal States, where that

oceanography determines the sensitivity of the ecosystem to nutrient enrichment. The geographic extent

of the shelf in which a State has jurisdiction is a useful place to begin classification. Here one should

consider whether the shelf is wide or narrow (e.g., mid-Atlantic versus Pacific Coast). The Texas coastal

shelf is very wide, with a gentle slope compared with much of the northern Gulf of Mexico. The

steepness of the slope is another useful factor, as it may influence bottom sediment stability and
Nutrient Criteria—Estuarine and Coastal Waters 3-13

upwelling. The degree of bottom roughness or sculpture may influence vertical mixing, which may in

turn influence water column stability and depth of the euphotic zone versus mixing depth. 

Nongeomorphic Classification

Walsh characterized the world’s continental shelves on the basis of their location, major rivers, and rates

of primary production, and included some U.S. coastal waters for comparison. The shelf proper is where

oceanic and estuarine boundaries often intermingle. At the shelf edge, cold, nutrient-rich water and

associated materials intrude onto the shelf. There, exchanges often are rapid, promoting conditions

favorable for higher fertility than in the open ocean. Higher primary productivity is the main reason why

approximately 90% of the world’s fish catch is harvested on the continental shelves versus the open sea

(Alongi 1998).

Water Quality Trend Detection on the Shelf

Physical mixing and advective processes may add considerable variability to water column measures. 

Therefore, it is important to consider detection of trends in nutrient concentrations and measures such as

chlorophyll a based on comparisons at a reference salinity (e.g., 30 psu). Otherwise, classification

schemes may incur extraneous variability. A common approach is to use “mixing diagrams” to compare

measured changes in an ambient constituent among sampling periods. At mid- and higher latitudes,

winter measures of DIN and DIP may provide insight into long-term trends of changes in nutrient

concentrations available to drive the spring bloom. At low latitudes winter values will likely have less

applicability, as primary production has a smaller seasonal signal.

Presence of Large Rivers

Although large rivers are included in Walsh’s characterization of shelf systems, it seems useful to

distinguish shelf areas based primarily on large rivers, such as the Mississippi River in the Gulf of

Mexico and the Columbia River off the Washington-Oregon coast. Large rivers on the shelf dominate

local ecological relationships.

Hydrographic Features

Vertical salinity differences tend to decrease toward the open ocean boundary. The principal reason is

summertime thermal stratification. The thermocline tends to be deeper toward the open sea margin,

except where buoyancy effects are associated with large rivers that flow onto the shelf. Coastal waters

contain a variety of biotic communities, including a diverse assemblage of macroepifauna and -infauna,

kelp forests, coral reefs, bottom and pelagic fishes, marine mammals, and seabirds. The relationship of

these communities to physical-ordering factors can assist in classification.

Temperate and subtropical coastal waters also experience a seasonal sea-level fluctuation, whereby

summer levels rise approximately 0.2 m by upper-ocean heat expansion, producing what is known as

thermosteric effects (Pattulo et al. 1955, Bell and Goring 1998). This nontidal process operates in

conjunction with other factors affecting apparent mean sea level (e.g., near and far-field wind effects and

barometric pressure). Depending on local conditions, water levels overlying the continental shelf and in

estuaries can rise from 0.1 to 0.2 m. Such a rise may seem nominal but can have a significant impact in
3-14 Nutrient Criteria—Estuarine and Coastal Waters

wetlands and other low-lying areas that potentially exchange nutrients and suspended sediments with the

coastal ocean.

Coastal ocean waters range from quite cold (e.g., Gulf of Maine) to quite warm (e.g., Gulf of Mexico). 

Where large rivers enter coastal waters, such as the Mississippi River Plume and Chesapeake Bay Plume,

visual discoloration can be observed because suspended material from land runoff and relatively high

plankton concentrations contrast with the predominantly blue color of the open ocean. The Columbia

River and the Mississippi River form an “estuary” mostly at sea, as very little of the diluted seawater is

bounded by land.

Physical gradients are dynamic and change at multiple scales. Seasonal or wet and dry periods

frequently differ depending on the various shelf gradients associated with estuarine, riverine, and

ocean/shelf break processes. Regional geomorphology and physical mixing processes play a pivotal role

in energy flow and material cycles. For example, the Loop Current in the eastern Gulf of Mexico may

show seasonal reversals and vary seasonally in its penetration onto the shelf. Along-shore drift inside the

north-flowing Gulf Stream off the Mid-Atlantic Bight tends to transport materials southward toward the

North Carolina coast. Further south, the Gulf Stream forms a seaward boundary that tends to

significantly isolate in-shore waters from those beyond the shelf break. Local current maps are available

from the National Ocean Service of NOAA (www.noaa.gov; then click on nos).

Many different types of boundaries or fronts occur in coastal seas, but no formal classification exists. 

Alongi (1998) lists five categories:

• Shelf-sea (tidal) fronts

• Estuarine fronts or plumes

• Shelf-break fronts

• Upwelling fronts

• Island wakes and fronts caused by other land features

Fronts provide increased physical stability at local scales, which may positively influence primary

production and energy flow to higher trophic levels (see Chapter 2).

Habitat/Community Differences

Presence of Mangrove/Seagrass and Coral Communities

Along the southeastern Florida Atlantic coast exists a combination of mangrove, seagrass, and coral reef

ecosystems. In some localities, each community type may dominate the others, but often they co-occur. 

Seagrass communities may dominate certain shelf areas along the west coast of Florida (e.g., Big Bend

region). The Flower Gardens, a disjunct coral community, exist off the southern coast of Texas. Alongi

(1998) devotes chapters to coral reefs and mangrove ecosystems including factors regulating primary

productivity (e.g., N and P). The role of N and P enrichment versus grazing in coral reef ecosystems is

still strongly debated in the scientific literature (e.g., Miller et al. 1999). A paper by Chen and Twilley

(1999) discusses soil nutrient relationships and productivity in a Florida Everglades mangrove ecosystem

along an estuarine gradient (see the references cited above for the most recent perspective). These
Nutrient Criteria—Estuarine and Coastal Waters 3-15

distinctive ecosystems provide a basis for local coastal waters classification. Mangrove communities

also occur along the lower Texas coast, and seagrasses are a dominant community in Laguna Madre, TX.

Presence of Seaweed

Seaweeds are common algal communities in rocky intertidal zones (e.g., Fucus spp.), attaching

themselves by means of a holdfast. Seaweeds belong to three marine algal classes: Chlorophyceae

(green algae), Rhodophyceae (red algae), and Phaeophyceae (brown algae). The kelps (Laminariales),

members of the brown algae, live subtidally but in relatively shallow waters and can form large forests

along the cooler north Atlantic and Pacific coasts. These communities also may occur in the higher

salinity reaches of estuaries. Alongi (1998) provides a discussion of primary production, factors limiting

growth, nutrient cycling, and grazing in these communities.
Nutrient Criteria—Estuarine and Coastal Waters 4-1

Causal and Response Variables

Field and Laboratory Methods

Nutrient Enrichment and Ammonia Toxicity

CHAPTER 4

Variables and Measurement

Methods To Assess and Monitor 

Estuarine/Marine Eutrophic Conditions

4.1 INTRODUCTION

This chapter provides an overview of several measurable trophic state variables that can be used to

establish nutrient criteria for estuaries and nearshore coastal waters. Trophic state variables are those

variables that can be used to evaluate or predict the trophic status or degree of nutrient enrichment of

estuaries and nearshore coastal waters, especially when compared with reference conditions. The primary

variables include two causal variables (TN and TP) and two response variables including a measure of

algal biomass (e.g., chlorophyll a for phytoplankton or macroalgal biomass (AFDW) and water clarity,

e.g., Secchi depth or electronic photometer), and the addition of dissolved oxygen, as appropriate. These

variables are relevant at the national scale to practically all estuaries and are potentially relevant to

nearshore coastal waters.

Several variables are important indicators of nutrient overenrichment for a large number of estuaries, but

in many cases the data and supporting science are inadequate for most systems (e.g., algal species

composition). Important secondary variables include seagrass and estuarine submerged aquatic

vegetation (SAV) distribution and abundance, macroinfaunal community structure, phytoplankton

species composition, and organic carbon concentrations, respectively. Seagrasses and SAV typically

provide important shallow water habitat information, and hypoxia/anoxia are measures of loss of bottom

habitat often associated with deeper waters. Organic carbon (total, particulate, and dissolved) is also

included as a secondary variable because this variable is consistent with Nixon’s (1995) definition of

eutrophication. Changes in benthic macroinfaunal community structure often correlate with organic

carbon enrichment and degree of hypoxia and anoxia (Diaz and Rosenberg 1995). The importance of

algal species composition has implications for food webs (Roelke 2000). These variables are discussed

in Chapter 2. 

As indicated in Chapter 2, the concentration of the primary nutrient variables may not correlate well with

one or more response variables in estuaries, especially hypoxia or anoxia and measures of phytoplankton

biomass. In this case, predictive relationships should be attempted with nutrient loads using first

empirical regression models or other statistical approaches if necessary to account for ecosystem-based

nonlinearities. Application of mechanistic computer models is another approach (see Chapter 9). 

Interpretation of nutrient enrichment indicators, especially for estuaries, is complicated by the interaction

with measures of mixing and flushing as discussed in Chapters 2 and 3. Salinity gradients are associated

with flushing but also play an important role in the type of biological communities exposed to nutrient
4-2 Nutrient Criteria—Estuarine and Coastal Waters

enrichment. These physical considerations must always play a part in nutrient enrichment predictions

including establishment of reference conditions as discussed in Chapters 2 and 6.

4.2 CAUSAL AND RESPONSE INDICATOR VARIABLES

Nutrients as Causal Variables

Nitrogen

Nitrogen is one of the most important limiting nutrients of autotrophic assemblages (e.g., phytoplankton

and periphyton) incorporated into estuarine and nearshore coastal marine bioassessments. In those

estuaries where N has been demonstrated to limit algal biomass production, it typically does so at higher

salinities along the salinity gradient (Chapter 2). Most research has focused on the role of inorganic-N as

a stimulant to algal biomass production (Stepanauskas et al. 1999). However, about 70% of the dissolved

N transported by rivers worldwide (1012g yr-1 ) is dissolved organic N (DON) (Meybeck 1982). In

contrast to P, control of N sources is more difficult because diffuse gaseous sources of N (N2) can be

assimilated directly from the atmosphere by N fixation, a process conducted by a variety of bacteria and

cyanobacteria (blue-green algae). Also, dissolved inorganic N forms, especially nitrite and nitrate, are

highly soluble and do not precipitate easily or sediment out when freshwater enters the brackish zone of

estuaries as inorganic P is likely to do.

Total N measured as a water quality indicator consists of organic and inorganic forms. Although some

dissolved organic N may be used for algal growth, especially if remineralized by bacterioplankton

(Carlson and Graneli 1993; Seitzinger and Sanders 1999), it and particulate organic forms participate in

algal biomass production through recycling processes (Chapter 2). In systems with hypoxic or anoxic

conditions, the rate of decomposition is reduced. Although still an open question, apparently relatively

little of the DON is directly utilized by phytoplankton, except for urea and free amino acids (Antia et al.

1991; Paerl et al. 1999). Dissolved organic N in rainwater (synthetic addition of urea and other

constituents in bioassays) was shown experimentally to stimulate bacterioplankton and phytoplankton

growth; however, the DON resulted in the dominance of diatoms and dinoflagellates whereas

ammonium-N stimulated production more of small monads (Seitzinger and Sanders 1999). Further work

is required to test whether this response is widely applicable. Thus, the source of DON can influence the

degree of DON utilization by the microbial community. Inorganic N consists of ammonia, nitrite, and

nitrate N. Ammonia N is a primary product of microbial degradation of organic N, and, if not used

directly by autotrophic algae and vascular macrophytes and microbial heterotrophs for growth, it may be

oxidized through nitrification to nitrite and nitrate. Varying proportions of organic N may be relatively

refractive and contribute very little to N overenrichment problems. However, the readily recyclable

component may contribute to N enrichment problems locally and further seaward. Some experimental or

model analysis (e.g., box model) of the utilization of DON and in some cases particulate organic N for

each coastal system is usually warranted.

In estuaries, N concentrations, especially the inorganic forms, typically vary widely seasonally,

interannually, and along salinity gradients. In temperate river-dominated estuaries, nitrate concentrations

may reach very high concentrations (e.g., >100 μM) in tidal fresh to brackish reaches (see Appendix G;
Nutrient Criteria—Estuarine and Coastal Waters 4-3

Neilson and Cronin 1981) due to wash-off associated with various land use activities including point and

nonpoint sources (e.g., agricultural cropland). By late spring to early summer, the nitrate concentration

may be below analytical detection limits. Nitrite concentrations seldom reach high levels in surface

waters due to plant utilization and conversion to nitrate through nitrification. The principal bacteria

genera that mediate nitrification include Nitrosomonas, but species of Nitrosococcus, Nitrobacter, and

Nitrospina are also important (Sharma and Ahlert 1977, Watson et al. 1981). If dissolved oxygen is

limiting nitrification, then nitrite may accumulate (Helder and de Vries 1983). Ammonia concentrations

in open estuarine and nearshore coastal waters located away from point sources typically vary from

below detection limits to approximately 1.0 to 5 μM, depending on growing season and rates of organic

N decomposition. Much higher values may occur for relatively short periods. The ionized form of

ammonia/ammonium is the most abundant reduced form and represents approximately 97% of the total

(Sillen and Martell 1964). The equilibration between the ionized and un-ionized fractions is controlled

by temperature, salinity, and pH, resulting in a range of un-ionized ammonia of 1% to 5% of the total at

typical salinities, pH, and temperature (Emerson et al. 1975). Ammonia may be toxic to marine larvae,

not just a stimulus to algal growth. Unionized ammonia concentrations in the range of 1.0:M

approximate those that are known to be toxic to marine larvae, especially molluscs (U.S. EPA 1989). 

Denitrification may remove from a few to approximately 50% of the TN load entering temperate

estuaries annually (Seitzinger 1988, Cornwell et al. 1999) depending largely on residence time of the

water, sediment biogeochemical conditions (macroinfauna present to maintain irrigation, oxic conditions

in the overlying bottom water), and water column depth. This process helps to modulate extreme DIN

concentrations (Chapter 2). Typical values for dissolved inorganic N (DIN) and a few TN concentrations

in estuaries and coastal nearshore waters are presented in Appendix G as a basis to help establish

expectations for various coastal systems. It should be noted that N concentrations vary widely in space

and time and the values in Appendix G are only intended to be rough guides. Specifics of analytical

techniques to measure the various forms of N are included at the end of this chapter (Field Sampling and

Laboratory Analytical Methods).

In open coastal waters of the North Atlantic Ocean at temperate latitudes, there is a typical seasonal

progression in DIN and DIP concentrations associated with phytoplankton blooms. The spring bloom

reduces these inorganic forms while phytoplankton biomass accumulates. This progression begins at

lower latitudes and moves to higher latitudes. The spring bloom typically crashes in late spring, and

summer biomass levels often are nutrient limited. Often a small bloom occurs in the fall following the

fall thermocline breakdown that allows mixing and replenishment of nutrients from deeper waters into

the upper surface layers, where a short burst of production occurs before light becomes limiting. 

Accumulation of deepwater nutrients during the winter has been used to assess the potential for springsummer overenrichment in coastal seas based on trends in “salinity-nutrient mixing diagrams” (European

Union Northern Marine Eutrophication Criteria Program, Ulrich Claussen, Germany, personal

communication). Seasonal nutrient patterns in estuaries are quite variable. In some estuarine systems, a

winter buildup of N and P has been observed (e.g., Patuxent River Estuary), especially when freshwater

flows remained low and point sources dominated the nutrient supply (e.g., Flemer et al. 1970). Mixing

diagrams also help interpret nutrient behavior in estuaries; however, some precautions are important to

recognize (e.g., see Sharp et al. 1986).
4-4 Nutrient Criteria—Estuarine and Coastal Waters

At the interface between fresh and marine waters, a process occurs that results in an apparent increase in

the ionized ammonia concentration. This process is apparently driven by the increased electrolyte

solution of the salts, which has a significant impact on the production and nitrification process, thus

yielding higher ionized ammonia levels (Rysgaard et al. 1999). Ionized ammonia adsorption to particles

was decreased, especially in the 0 to 10% salinity range, as were the nitrification and denitrification

processes. Further evaluation showed that the reduction in nitrification and denitrification processes was

due not only to the displacement of bacteria and ionized ammonia from particles, but also to decreased

bacterial activity. The projections from these studies were that ionized ammonia would be produced at a

rate of 1 μM/g of sediment in the water. The changes in N dynamics that affect adsorption of suspended

solids may need to be included when considering acceptable levels in fresh water sources to estuaries.

Phosphorus

Phosphorus is an important plant nutrient that may limit algal biomass production in tidal fresh to

brackish zones of estuaries and some subtemperate to tropical marine coastal systems (Chapter 2). There

are no common stable gaseous forms of phosphorus, so the phosphorus cycle is endogenic, without an

atmospheric component (Manahan 1997). The main natural reservoirs of phosphorus are poorly soluble

minerals (e.g., hydroxyapatite) in the geosphere. Erosion of these materials from terrestrial sources and

their transport to the sea are important sources of new phosphorus in seawater. The phosphorus entering

the sea is mostly orthophosphate, PO4

-3 (Kennish 1989). In previous decades, prior to widespread

phosphate bans in detergents, estuaries received a considerable portion of P from detergents. The ban

resulted for many estuarine systems in an elevated DIN:DIP ratio. In estuaries and nearshore coastal

waters, phosphorus is present in dissolved inorganic form as well as dissolved and particulate organic

form. Some fraction of P may be strongly embedded in a mineral matrix, and this renders that fraction

relatively inert to biological utilization. For this reason, often measures of TP may represent some

component that is not biologically available and managers should consider this in developing P criteria.

Plants directly take up the phosphates as essential nutrients during photosynthesis. Some algae have the

capability to break down dissolved organic P (DOP) with alkaline phosphatase (algal and free

phosphatases) and utilize the phosphate as inorganic phosphate (Huang and Hong 1999). Alkaline

phosphatase apparently is located on phytoplankton cell membranes, which makes it difficult to

determine whether the uptake is direct for DOP or the DOP undergoes enzymatic hydrolysis on the cell

membrane. Malone et al. (1996) suggested by inference that Chesapeake Bay phytoplankton may utilize

organic sources of P, in part, because the DIN:DIP thresholds approach 160, which is considerably

greater than the N:P ratio reported by Redfield et al. (1963). Orthophosphates are typically preferred by

autotrophic phytoplankton, although some assimilation of organic phosphorus may occur, especially

during periods of P deficiencies (Boney 1975). When plants die, or are eaten, the organic phosphorus is

rapidly converted to orthophosphates through the action of phosphorylases within fecal material,

phosphatases in the plant cells, and finally by bacteria (Riley and Chester 1971). 

To summarize, phosphorus occurs in natural waters and in wastewaters almost solely as phosphates. 

These are classified as orthophosphates, condensed phosphates, and organically bound phosphates

(common analytes are total phosphorus [TP] and dissolved or particulate organic phosphorus [DOP,

POP]). These compounds may be soluble, in particulates or detritus, or incorporated as organic P in
Nutrient Criteria—Estuarine and Coastal Waters 4-5

organisms. Phosphorus is essential to the growth of organisms and can limit phytoplankton biomass

production, which is most commonly observed in freshwater systems (Hecky and Kilham 1988) and

some estuaries and coastal marine systems (Chapter 2). In instances where phosphate is limiting, the

discharge of raw or untreated wastewater, agricultural drainage, or certain industrial wastes may

stimulate the growth of algae. Appendix G provides examples of P concentrations in several forms.

Silica

Silica, as an important algal nutrient, has received much less attention in estuarine nutrient

overenrichment studies than N and P based on the limited volume of literature citations (e.g., see Malone

et al. 1996) and recent reviews of estuarine eutrophication (Chapter 1). Silica limitation of diatom

production, a major algal group that requires Si (and silicoflagellates), often is a measure of N or P

overenrichment (D’Elia et al. 1983; Conley and Malone 1992). Dissolved Si is a product of weathering

and erosion of rocks on land with subsequent transport to the sea (Conley and Malone 1992). Because Si

has essentially no human sources, except possibly from erodible soils under human influence, it is not a

strong candidate for regulation. In some parts of the ocean, organisms (such as diatoms and radiolarians)

abound that have produced skeletons of a noncrystalline form of hydrated silica-opal. As these skeletons

settle to the sea floor they slowly dissolve, releasing silica. Officer and Ryther (1980) predicted that

increases in N and P to estuaries and coastal waters from human activities, coupled with the reduction in

silicates to the sea from construction of artificial lakes, would alter the N:Si and P:Si ratios. These

alterations were postulated to alter phytoplankton populations to reduce the relative abundance of

diatoms and enhance the relative abundance of flagellates. Egge and Aksnes (1992) showed that diatoms

always numerically dominated the phytoplankton community when concentrations of silica were in

excess of 2.2 μM. Dominance by diatoms ceased or became more variable when concentrations of Si

were less than this value.

Ryther and Officer (1981) reinterpreted the relationship of N pollution in Long Island Inlets during the

1950s. Nitrogen may have limited the nuisance Nannochloris, blooms but they hypothesized that the

bloom persisted because diatoms had been eliminated by Si depletion. Also, the degree of Si limitation

of spring diatom blooms in Chesapeake Bay that fuel summer anoxia has direct ecological implications

(Conley and Malone 1992, Malone et al. 1996). Freshwater sources of Si dominate estuarine supplies

(Fisher et al. 1988). Typically, Si limitation can be potentially deduced from ambient ratios relative to

the nutrient-sufficient N:Si:P biomass ratios of 16:16:1 (Redfield et al. 1963; Conley et al. 1993). In

Chesapeake Bay, the dissolved Si:DIP ratio often approximates 100-300 (Malone et al. 1996), suggesting

strong Si limitation. Significant increases in Mississippi River N and P concentrations and loading and

decreases in silicate have occurred during the 20th century (Rabalais et al. 1996). The increased P

loading and associated increased diatom production and eventual burial in river sediments, as predicted

by Officer and Ryther (1980), has resulted in a reduced Si supply to the coastal environment. The

consequence is that diatom production, generallly a preferred phytoplankton group to support higher

trophic levels, is now more Si limited than in previous decades. The N:P:Si ratios on coastal Louisiana

and Texas now suggest the possibility of a joint nutrient limitation of phytoplankton production. 
4-6 Nutrient Criteria—Estuarine and Coastal Waters

Silica concentrations for the Coastal Texas/Louisiana coast averaged approximately 5.3 μM in the late

1980s but averaged about 9.0 μM during the early 1960s. Silicate concentrations in the Chesapeake and

Delaware Bays and the Hudson River Estuary ranged from about 90 to near detection levels, 30 to near

detection limits, and 30 to 3 μM, respectively (Fisher et al. 1988). Eyre and Balls (1999) reported that Si

was less likely to limit diatom production in tropical estuaries than in temperate ones because

concentrations tend to be much higher in tropical estuaries.

The role of silica may be more important to diatom species composition and food quality as future

research may document. More attention in the future should be given to the measurement and assessment

of the role of Si in estuarine and nearshore coastal primary productivity and food web dynamics and as a

basis for controlling co-limiting N and/or P.

Response Variables

Chlorophyll a and Macroalgal Biomass

Chlorophyll a is the molecule mediating photosynthesis in most all green plants (except prochlorophytes,

which contain divinyl chlorophyll), including phytoplankton; it is relatively easy to measure either

spectrophotometrically or by fluorescence and is commonly used to indicate phytoplankton biomass. 

However, the amount of chlorophyll per cell can vary widely. Conversion factors from weight of

chlorophyll to weight of carbon (a desired biomass unit) can vary by a factor of 10. Adaptation to light

levels is the primary reason for observed variability; photoadaptation can cause the chlorophyll per cell

to vary widely. The technology for measuring chlorophyll has greatly improved over the decades. The

Welschmeyer (1994) fluorometric analysis reduces the interference due to chlorophyll b and

phaeopigments. The HPLC procedure is capable of detecting and quantifying various pigments

characteristic of different algal groups (e.g., diatoms, cyanophyta, chlorophyta, and dinoflagellates)

(Jeffery et al. 1997).

Rapid proliferation or blooming of phytoplankton, as reflected in chlorophyll a measurements, occurs

throughout the ocean but is most often associated with temperate coastal and estuarine waters and at

higher latitudes. In winter months, growth of phytoplankton populations is generally minimal because of

insufficient light and also because a turbulent and unstable upper water column carries the phytoplankton

cells below the euphotic zone (where light is not sufficient) before they can divide.

Chlorophyll a concentrations vary widely as a function of nutrient supply, water column stability,

euphotic zone depth (light availability), sinking, grazing, disease organisms (e.g., viruses), and

flushing/mixing (Chapter 2). Values in excess of 12 to 15 μg/L are likely to cause severe shading of

seagrasses (Kelley in press). Concentrations in estuaries during summer optimum growing conditions

may exceed 50 to 80 μg/L when nutrient loading is high (Monbet 1992). Summer values in the range of

20 to 40 μg/L are frequently observed in enriched estuaries. In contrast, concentrations in overenriched

temperate U.S. estuaries during the winter may decrease to 1 to 5 μg/L Nearshore coastal areas removed

from high nutrient loads may experience chlorophyll concentrations in the range of approximately 1 to 3 

μg/L (Appendix G). Very high values may occur during the summer under conditions of high levels of
Nutrient Criteria—Estuarine and Coastal Waters 4-7

nutrient enrichment (e.g., the Mississippi River Plume on the Texas/Louisiana Shelf [Rabalais et al.

1996]).

Macroalgal biomass, especially benthic unattached forms (i.e., Ulva spp.), often becomes abundant in

relatively shallow estuaries that experience nutrient overenrichment. In estuaries that receive most of

their nutrient load from groundwater (e.g., Waquoit Bay, Cape Cod, MA; see Chapter 2) benthic

macroalgae may shade out seagrasses. Continued enrichment typically leads to reduction of macroalgae

as phytoplankton predominate in the water column. Macroalgae are difficult to adequately sample for

chlorophyll a, and thick mats often contain sheets of algal material that has begun to degrade. The most

common method to sample benthic macroalgae is to collect samples and express the biomass on a dry

weight basis.

Measures of Water Clarity

Light Attenuation Coefficient

The Secchi disc has been a mainstay as a tool in estimating water clarity; however, this simple and

inexpensive tool does not provide all of the information required to distinguish the light attenuation

effects of living phytoplankton pigments (i.e., traditionally estimated by chlorophyll a) from other factors

(e.g., inorganic suspended sediments, organic nonchlorophyll-based detritus, and humic-like materials)

that reduce water clarity. EPA’s Chesapeake Bay Program (Chapter 2) has developed an analytical

approach that partitions the effect of chlorophyll a from total suspended solids that contribute to

reduction in water clarity. This approach has been used successfully in estimating the combined factor

contribution to light attenuation over submerged aquatic vegetation beds (Dennison et al. 1993). In

turbid coastal waters, the analyst should be aware of lower values for the constant 1.7 to estimate the

light attenuation coefficient (see Giesen et al. 1990 and references in Chapter 2). More precise estimates

of the light attenuation coefficient can be made with electronic submersible light meters including PAR

meters (photosynthetic active radiation) and submersible spectral radiometers. These meters are now in

widespread use, and their use should be encouraged because they give a direct measure of light

attenuation, especially in shallow water where depth may limit use of the Secchi disc. 

Attenuation of light in the sea in nonalgal bloom areas is determined principally by the amount of

suspended matter present, but in estuaries and nearshore coastal waters, color from humic-like materials

may significantly compete with particulate material in light attenuation. In moderately turbid coastal

waters, 1% of the surface visible light energy may penetrate to a depth of only 10 to 20 m, but in shallow

estuaries depths often are from 10 cm to 3 m or so. There typically is a strong seasonal variability in

water clarity in temperate estuaries between the active growing season and the winter, and in

subtemperate to tropical estuaries water clarity often is a function of the wet season. In the Atlantic

temperate open coastal areas with the coming of spring, the depth of the euphotic zone often increases

and the depth of the mixed layer decreases because of the development of the seasonal thermocline. This

allows a spring bloom to develop. The thermocline tends to confine the algal cells to the euphotic zone,

which becomes rich with nutrients as a result of winter mixing. In estuaries, the pycnocline may also

have this effect. In partially mixed estuaries where light is adequate at depth, diatoms may grow below

the pycnocline (Malone et al. 1996). If the necessary growth-promoting factors are also present,
4-8 Nutrient Criteria—Estuarine and Coastal Waters

conditions are optimal for proliferation of phytoplankton from seed stock, which may be either the

plankton cells themselves or their resting stages (Riley and Chester 1971). 

Secchi Depth

The Secchi disc is a useful tool to estimate water clarity (Holmes 1970). Secchi disc measurements often

have a longer historical record than electronic measurements, which facilitates assessment of trends in

water clarity. Secchi depth measurements are obtained with a 40 cm plastic or metal Secchi disk that is

either white or is divided into black and white quadrants on a nonstretchable line that is calibrated in

decimeters. The disc should be weighted to maintain a level position, especially under strong current

conditions. The disk is lowered into the water until it disappears from view and the depth is recorded. 

The disk is then slowly raised to the point where it reappears and the depth is recorded again. The mean

of these two measurements is the Secchi depth. Observations are made from the shady side of the vessel

to reduce problems of glare; however, when a small boat is used for field work a “viewing tube” allows

readings under full sunlight conditions. Measurement should be made without sunglasses.

Dissolved Oxygen

Dissolved oxygen (DO) is an integrative measure of ecosystem health and habitat function. As a firstorder estimate, the percent saturation of surface and bottom waters is an index of the

production/respiration ratio. Dissolved oxygen in bottom waters serves as a measure of habitat

availability for benthic animals and pelagic animals that feed on the bottom. EPA has developed

saltwater DO criteria for coastal waters between Cape Cod and Cape Hatteras (see

www.epa.gov/ost/standards/dissolved). Profiles of DO are indicative of oxygen depletion conditions

such as hypoxia and anoxia. Lack of oxygen in bottom waters causes sediment to release dissolved

nutrients including orthophosphorus, ammonia, and in addition, toxic hydrogen sulfide. 

Carbon Compounds

Organic matter content is typically measured as total organic carbon (TOC) and dissolved organic carbon

and is an essential component of the carbon cycle. The rate of organic carbon production and

decomposition and the resulting microbial biomass are at the heart of the eutrophication problem. 

Evaluation of the carbon-containing compounds in an aquatic ecosystem can indicate its organic

character. The larger the carbon or organic content, the greater the growth of microorganisms that can

contribute to the depletion of oxygen supplies. TOC is a more convenient and direct expression of

organic carbon content than are the biochemical oxygen demand (BOD), assimilable organic carbon

(AOC), or chemical oxygen demand (COD) methods. TOC is independent of the oxidation state of the

organic matter and does not measure other organically bound elements, such as N and hydrogen, or

inorganics that can contribute to the oxygen demand measured by BOD and COD. In spite of its

versatility, TOC does not provide the same kind of information as BOD, AOC, or COD, and should not

be used to replace these methods.

At the surface of the sea, the concentrations of particulate and dissolved organic carbon range up to 12.5

μM and between 75 and 150 μM, respectively. In coastal environments, concentrations of dissolved and

particulate organic carbon are greater by factors of ~7-fold. Concentrations of dissolved and particulate
Nutrient Criteria—Estuarine and Coastal Waters 4-9

organic carbon in surface waters are equivalent to 150 to 1,800 μg C/L (Millero 1996). Organic carbon

represents approximately 50% of the dissolved and particulate organic material in seawater (Millero

1996). However, the major form of carbon in seawater is associated with inorganic carbonate systems. 

Benthic Macroinfauna

Benthic macroinfauna are an important biological component of estuarine and nearshore coastal marine

ecosystems. These communities contribute to benthic food webs, contribute to nutrient cycling and

system productivity through benthic-pelagic coupling of nutrient recycling, help stabilize bottom habitats,

and contribute to marine biodiversity. Benthic infaunal communities are quite diverse within an estuary

or coastal region. Diversity is a function of salinity, with higher diversities associated with higher

salinities (Carriker 1967). Sediment irrigation provided by benthic infauna enhances denitrification by

increasing the flux of ammonium into oxic microenvironments where nitrification can occur and the flux

of nitrite and nitrate into the anoxic sediment zone where denitrification becomes possible (Chapter 2). 

4.3 FIELD SAMPLING AND LABORATORY ANALYTICAL METHODS

The following sections provide additional information on field sampling and laboratory methods for

selected variables. A list of suggested methodologies for analysis of biochemical parameters is provided

in Table 4-1. These methods have been summarized from nationally or regionally recognized reference

compendiums (APHA 1998, ASTM 1976, U.S. EPA 1979, Spotte 1992) and provide acceptable methods

for determining the concentrations of nutrients as well as acceptable methods for measuring the effects of

those nutrients in estuarine and marine waters.

Field Sampling Methods

Nutrients, Hydrography, and Sediments

Physiochemical profiles should be recorded for each field sampling station. Important parameters to be

measured include water temperature, pH, dissolved oxygen, salinity, light attenuation, surface radiation,

and total depth. Generally, a multiparameter water quality instrument CTD is used. Sampling depth will

vary depending on specific objectives; however, enough vertical depth reading should be taken to

characterize the physical structure of the water column. For example, CDT measurements might be taken

at frequent intervals in the vicinity of the pycnocline (e.g., every 0.1 m in highly stratified estuaries). 

Overall current dynamics can be mapped with oceanographic tools such as current meters, drift cards,

and acoustic Doppler sounders.

Field sampling of discrete water samples for laboratory analysis can be performed using standard

nonmetallic plastic water bottles. Samples are drawn into prelabeled bottles and fixatives are applied as

appropriate to the subsequent analysis. Nutrient and organics samples are stored on ice until reaching a

shoreside sample handling location. Nutrient samples are filtered using graduated syringes and then

frozen. Samples for total TN and TP are filtered or unfiltered as appropriate, and 20 mL of sample is

frozen for analysis. See Chapter 5 for additional sampling protocols.
4-10 Nutrient Criteria—Estuarine and Coastal Waters

Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine

environments (* = primary EPA preferred causal and response variables)

Eutrophication indicators Suggested methods Detection limit

or range

Comments References

 Field

 *Water clarity Secchi depth 0.1 m — EPA 903-R-96-006

 pH CTD probe 0.01 pH — —

 Dissolved oxygen CTD probe 0.02 mg DO/L or Winkller Azide Mod. —

 Salinity Salinometer 0.1 psu — —

 Light attenuation Sensor 0.05%@100%light e.g., LI-COR-LI-192S A

sensor

—

 Temperature CTD probe 0.1°C — —

 Laboratory analyses

 *Total phosphorus SM 4500P-E 0.3 μM Ascorbic acid method APHA 1998

 (including

 orthophosphate, SM 4500P-E 0.32 μM Auto. persulfate method APHA 1998

 POP, and DOP) EPA 365.2 — — EPA 600/4-79-020

CBP IV.D.2 0.03 μM Auto. persulfate method EPA 903-R-96-006

 Dissolved

 orthophosphate CBP IV.D.3 0.02 μM Ascorbic acid method EPA 903-R-96-006

 Particulate phosphorus CBP IV.D.4 0.04 μM Ascorbic acid method EPA 903-R-96-006

 *Total N, incl. DON, SM 4500N-C 0.36 μM Persulfate method APHA 1998

 DIN, and PONa ASTM D3867 0.7-143 μM Persulfate method ASTM 1976

EPA — Persulfate method EPA 903-R-96-006

EPA-AERP18 — — EPA 600/4-87-026

CBP IV.D.8 1.9 μM Auto. persulfate method EPA 903-R-96-006

 Total Kjeldahl N SM 4500org-C with — Semi-micro-Kjeldahl method APHA 1998

SM 4500NH3-H 1.4-1429 μM Auto. phenate method APHA 1998

EPA 351.3/.1

(mod.) — Colorimetric/titration EPA 600/4-79-020

 Ammonia/ammonium SM 4500NH3-B/H 1.4-1429 μM Auto. phenate method APHA 1998

EPA 350.1 0.7-1429 μM Colorimetric phenate EPA 600/4-79-020

CBP IV.D.7 0.3 μM Auto. phenate method EPA 903-R-96-006

 Nitrate SM 4500NO3-F 35.7-714 μM Auto. cadmium reduction APHA 1998

EPA 353.2 — — EPA 600/4-79-020

 Nitrite SM 4500NO3-F 35.7-714 μM Auto. cadmium reduction APHA 1998

EPA 353.2 — — EPA 600/4-79-020

SM 4500NO2-B 0.7-71 35.7-714 μM Colorimetric method APHA 1998

EPA 354.1 — — EPA 600/4-79-020
Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine

environments (* = primary EPA preferred causal and response variables) (continued)

Eutrophication indicators Suggested methods Detection limit

or range

Comments References

Nutrient Criteria—Estuarine and Coastal Waters 4-11

CBP IV.D.5 0.01 μM Auto. colorimetric method EPA 903-R-96-006

 Nitrate + nitrite SM 4500NO3-F 35.7-714 μM Auto. cadmium reduction APHA 1998

EPA 353.2 — — EPA 600/4-79-020

EPA 4.1.4 0.7-143 μM Technicon autoanalyzer EPA 503/2-89/001

CBP IV.D.6 0.01 μM Auto. colorimetric method EPA 903-R-96-006

 Particulate N CBP IV.D.8.10 1.36 μM Filtration/combustion EPA 903-R-96-006

 Total organic carbon SM 5310TOC-D >0.1 mg C/L Wet oxidation method APHA 1998

SM 5310TOC-C >0.01 mg TOC/L Persulfate method APHA 1998

EPA 415.1 — — EPA 600/4-79-020

 Dissolved organic

 carbon SM 5310TOC-C >0.01 mg TOC/L Persulfate method APHA,1998

EPA 415.1 — — EPA 600/4-79-020

ASTM D2574-79 — — ASTM 1976

CBP IV.D.10 0.5 mg/L Catalytic combustion EPA 903-R-96-006

 Particulate carbon CBP.IV.D.9 0.097 mg/L Filtration/combustion EPA 903-R-96-006

 Total silicates SM 4500SiO2-D 0.33-0.83 μM Heteropoly blue method APHA 1998

ASTM D859-68 — — ASTM 1976

CBP-IV-15 0.17-23.3 μM EPA 903-R-96-006

EPA 370.1 — — EPA 600/4-79-020

CBP IV.D.15 0.22 μM Molybdosilicate method EPA 903-R-96-006

 Total suspended solids SM 2540-D 2-20,000 mg/L Dried at 103-105°C APHA 1998

CBC IV.D.13 2.0 mg/L Filtration/heat EPA 903-R-96-006

 Total volatile solids SM 2540-E — — APHA 1998

Estuarine — — EPA 430/9-86-004

 BOD SM 5210-B — 5-day method APHA 1998

EPA 405.1 — — EPA 600/4-79-020

CBP IV.D.11 — 5-day method EPA 903-R-96-006

 COD SM 5220-D — — APHA 1998

EPA 410.4 — — EPA 600/4-79-020

 Biological measures

 Phytoplankton biomass — — — —

 Zooplankton biomass

 Chlorophyll a

b SM 10200-H 0.01 mg/M3 Fluorometric, HPLC, Spectro. APHA 1998
Table 4-1. Suggested methods for analyses and monitoring of eutrophic conditions of coastal and marine

environments (* = primary EPA preferred causal and response variables) (continued)

Eutrophication indicators Suggested methods Detection limit

or range

Comments References

4-12 Nutrient Criteria—Estuarine and Coastal Waters

EPA AERP12 — — EPA 600/4-87-026

ASTM D3731-79 Spectrophotometer ASTM 1976

CBP IV.D.12 1.0 μg/L Spectrophotometer EPA 903-R-96-006

 Phaeophytin SM 10200-H 0.01 mg/M3 Fluorometric, HPLC, Spectro. APHA 1998

EPA AERP12 — — EPA 600/4-87-026

ASTM D3731-79 — Spectrophotometer ASTM 1976

CBP IV.D.12 1.0 μg/L Spectrophotometer EPA 903-R-96-006

 Dinoflagellate density — — — —

 Diatom density — — — —

 Dinoflagellate/diatom — — — —

 Perennial plant density — — — —

 Ephemeral plant density — — — —

 Epiphytic growth — — — —

 Phytoplankton blooms — — — —

 Fish kills — — — —

a

 DON, dissolved organic N, DIN, dissolved inorganic N; PON, particulate organic N.

b

 Phytoplankton segments: The HPLC procedure is capable of detecting and quantifying various pigments characteristic of

different algal groups (e.g., diatoms, cyanophyta, chlorophyta, and dimoflagellates) (Jefferey et al. 1997).
Nutrient Criteria—Estuarine and Coastal Waters 4-13

Laboratory Analytical Methods

Detailed methods and references are given in Table 4-1. Some general considerations are presented in

the following sections.

Water Column Nutrients 

Nitrogen Compounds

Several methods have been used to determine the concentration of N species in the marine environment. 

Methods presented in this document are relatively easy to use, do not require extensive instrumentation,

provide detection limits below those expected in marine environments, and are in general use by many

investigators. The most common forms of N in eutrophication evaluation in order of decreasing

oxidation state are nitrate, nitrite, ammonia, and organic N. The sum of these is expressed as TN and is

not to be confused with total Kjeldahl N (TKN), which is the sum of organic N and ammonia. Total N

can be determined through oxidative digestion of all digestible N forms to nitrate, followed by

quantitation of the nitrate. Nitrite is an intermediate oxidation state of N, both in the oxidation of

ammonia to nitrate and in the reduction of nitrate. Such oxidation and reduction may occur in

wastewater treatment plants, water distribution systems, or natural waters. Ammonia is produced largely

by deamination of organic N-containing compounds and by hydrolysis of urea. The two major factors

that influence selection of the method to determine ammonia are concentration and presence/absence of

interferences (e.g., high concentrations of colored organic substances such as humic-like materials or

paper mill effluents).

Total N is measured by the persulfate method, which digests all organic and inorganic – containing

compounds. All N-containing materials (except nitrogen gas) are measured after sample digestion has

occurred. Various organizations have adjusted sample volume or automated the process and produced

different ranges of detection. The lowest detectable concentration is ~ 0.7 μM of TN. This is in the

range of the measured available N (0.7 to 5.0 μM TN) for studies performed off the continental shelf in

the North Atlantic from 1956 to 1958 (Kennish 1989). Kjeldahl N minus the ammonia concentration is

the surrogate measurement for all organic N-containing compounds.

Ammonia/ammonium is measured by the indophenol blue (= phenate) or specific ion electrode methods

after conversion of ammonia and ammonium to ammonia. This is done by raising the pH of the sample

above 11. This method has some essential features (e.g., minimal interference from waters highly stained

with humic materials and paper mill effluents); however, the level of detection is relatively high (e.g., 2.0

μM NH3-N) but adequate for ammonia-rich waters (Flemer et al. 1998). Ammonia electrodes do not

work directly in seawater. In the spectrophotometric methods, the ammonia is reduced to

monochloramine and then reacted with phenol to form a blue color. In the specific ion electrode method,

the ammonium is converted to ammonia using a strong basic solution and partial pressure of ammonia

gas (i.e., free ammonia) in solution, which is related to the dissolved ammonia concentrations by Henry’s

Law. 

Nitrates and nitrites are measured in combination using the cadmium reduction procedure of Wood et

al. (1967). This colorimetric method determines the concentration of these two materials after reaction
4-14 Nutrient Criteria—Estuarine and Coastal Waters

of nitrites to produce an azo dye, the color of which is proportional to the concentration of the combined

nitrates and nitrites. Total nitrate is determined by subtracting the concentration of nitrite from the

combination of the two. The process for measurement of nitrite produces the same azo dye as the

combined measure, but without the Cd reduction. The difference in these two measures is the nitrate

concentration. 

Phosphorus

The target detection limit for measurement of P in seawater is ~ 0.3 μM. The procedures for the

measurement of total particulate and dissolved P as well as orthophosphate in seawater provide detection

limits that are less than this value (U.S. EPA 1996). These procedures convert the phosphoruscontaining compounds to orthophosphate through the digestion of the sample with alkaline persulfate. 

This treatment is then reacted with ammonium molybdate and antimony potassium tartrate in acidic

solution to produce an intense blue complex with ascorbic acid. Interferences with elevated

concentrations of Si can be avoided by maintaining an acid concentration in the reagents and analyzing

the material at elevated temperatures of ~37°C. The resulting phosphomolybdic acid reduction produces

a purple-blue complex that is measured at 885 nm on a spectrophotometer. This method of measuring

reactive silicate is recommended in Millero (1996).

Silica 

The target detection limit for measurement of Si in seawater is ~0.7 μM. Producing pigmented

silicomolybdate complex by procedures contained in U.S. EPA (1996) provides adequate sensitivity after

the samples are filtered (0.45 μm GF/F filter) to remove interfering particles and turbidity, and after the

interferences of phosphates and arsenates are removed with oxalic acid. The resultant filtrate is treated

with a solution containing metol-sulfate (p-methyl-amino-phenol sulfate) to produce a blue color that is

evaluated more efficiently than the yellow color recommended for evaluation in U.S. EPA (1996), with a

spectrophotometer at 812 nm (Strickland and Parsons 1968). This method of measuring reactive silicate

is also recommended in Millero (1996).

Carbon

Total carbon consists of inorganic and organic forms that are in particulate and dissolved size classes. 

The distinction between total and organic carbon is based on acidifying samples to remove the inorganic

forms and filtering through 0.45 μm GF/F filters to remove the particulate forms. Total carbon is

measured by burning the sample to release the particles contained on the glass fiber filter. This converts

the carbon to CO2, which is then transported to a thermal conductivity detector for measurement. The

carbon left behind in the filtrate is catalytically combusted using a platinum catalyst at ~680°C that is

then transported to a nondispersive infrared detector. The EPA methods (U.S. EPA 1996) will provide

adequate detection of both dissolved and particulate carbon in the total and organic phases. The

difference in total carbon and organic carbon represents the inorganic fractions that are primarily CaCO3

shells.
Nutrient Criteria—Estuarine and Coastal Waters 4-15

Sediment Analyses

Bulk Sediment

Cores are collected from field sites to help determine the historical record and sedimentation rate. Short

cores, the upper 30 cm of the substrate, can be obtained with a HAPPS core, designed to collect a

relatively undisturbed core of surficial sediment (Kannerworff and Nicolaisen 1973) and used to profile

sedimentary particulate organic carbon and N. Carbon-N analyses follow the method of Hedges and

Stern (1984); samples for dissolved constituents in pore water are extracted either by whole-core

squeezing or by centrifugation (Devol et al. 1997, Brandes and Devol 1995, Lambourn et al. 1991). Deep

coring devices are used to collect continuous sediment core samples 2 to 3 m into the sediment bed. 

These deeper cores are used for analysis of 210Pb, carbon and N, sulfide, and biogenic silica in order to

determine burial rates of 210Pb and 210Ra. 

The sedimentation rate is estimated based on the change in activity of naturally occurring 210Pb

radionuclide produced at a constant rate from the decay of 210Ra, using the excess 210Pb inventory method

of Anderson et al. (1987). Excess 210Pb is determined from the difference between total 210Pb activity in

the sediment and the activity of the background 210Pb being produced from 210Ra. To collect samples for

measurement of 210Pb and 210Ra activity at depth with the sediment, cores are sectioned and each section

is then homogenized and placed in a precleaned 16 oz jar, with a small subsample removed and placed

into a glass vial for particulate C and N analysis (Evans-Hamilton, Inc. 1998).

The excess 210Pb inventory method yields accumulation rates (g/(cm2/yr)), which are converted to a

sedimentation rate (cm/yr) using the bulk sediment density g/cm3. For evaluation of seasonal trends, the

upper cm is subsampled at 0.25 cm intervals, and in 1 cm intervals below the first cm, following the

assumption that any seasonal storage of N or carbon would manifest almost entirely at the surface of the

sediment.

Pore Water Profiles

Pore water profiles of manganese, iron, nitrate, and oxygen demonstrate that oxidation of iron and

magnesium yields less energy than does oxidation of carbon by oxygen or nitrate. Consequently,

concentration peaks of these species are located below the depletion depths of oxygen and nitrate. In

anaerobic environments, after the supplies of oxygen, nitrate, manganese, and iron are exhausted, sulfate

reduction is the dominant mode of organic matter oxidation and nutrient remineralization. 

Sulfate reduction rate can be measured with the radiotracer method of Christensen et al. (1987). A

significant fraction of the oxygen flux may be consumed by the reoxidation of sulfide produced during

sulfate reduction (Canfield 1993).

Sediment traps are used to measure the quantity and composition of the flux of materials settling through

the water column to the sediment. There are four materials of interest: chlorophyll as an indicator of

planktonic algal remains, pheaopigments as an indicator of degraded plankton that has been consumed by

zooplankton, particulate organic carbon (POC), and particulate organic N (PON). Total sedimentation

rate is corrected for resuspension materials in order to derive the net flux to sediment. Samples are
4-16 Nutrient Criteria—Estuarine and Coastal Waters

collected by in situ benthic flux chambers, and measurements of oxygen, silicate, nitrate, ammonium,

phosphate, and N gas are made (Evans-Hamilton, Inc. 1998).

Determination of Primary Productivity

Primary productivity refers to the growth rate of the phytoplankton community and is commonly

measured using trace amounts of radioactive carbon (as bicarbonate) that label the photosynthetic

reaction. Additional variables are measured to support these data: biomass (as estimated by chlorophyll

a), incoming solar radiation, and nutrient concentrations at depth. Primary productivity, P, is defined as

P = μ × B

where μ is the specific growth rate (growth normalized per cell) and B is the biomass of the

phytoplankton population (amount of cells). These variables are > ‘compound’ = as they in turn depend

on other variables. Growth rate depends on light (solar radiation), dissolved nutrients in the water

column, and water temperature. The phytoplankton biomass is determined by the net result of growth

and loss (grazing, mixing, sinking) processes and reflects enrichment conditions.

To estimate primary productivity, samples are collected at varying depths corresponding to

predetermined light levels. Fresh samples at each light level are collected for analysis of chlorophyll a,

nutrients, and primary productivity in two sets of two clear bottles and one dark bottle; each set is filled

for ambient treatment and nutrient spike treatment. Nutrient spiking consists of adding an initial

concentration of 10 μM N (NH4CL) and 1 μM phosphorus (KH2PO4) to seawater. Nutrients are

monitored from additional samples collected and tested for nitrate, nitrite, ammonium, orthophosphate,

and silicate. Samples are inoculated with 14C-labeled sodium bicarbonate and, if appropriate, the nutrient

spike, and placed in a screened bag to simulate the light level from which they were collected. Samples

are incubated at in situ conditions for 24 hours and then transported to the laboratory for filtration using

glass fiber filter paper (Whatman GF/F, nominal pore size 0.7 μm or smaller pore size). The filters are

placed into vials containing EcoLume scintillation cocktail. The specific activity of the filtered

particulates is measured in a scintillation counter. Primary production is calculated as mg C/(m3/day)

using the basic equations found in Parsons et al. (1984) (Evans-Hamilton, Inc. 1998).

In productive coastal waters, measurements using the light and dark bottle technique with changes in

dissolved oxygen often can be used in place of the 14C method (Strickland and Parsons 1968). In some

cases, free water gas-based (e.g., DO) methods are possible to measure ecosystem metabolism (Odum

1956; Odum et al. 1959; Kemp and Boynton 1980).

Phytoplankton Species Composition

Samples collected from the field are analyzed to identify and enumerate autotrophic phytoplankton, as

well as heterotrophic dinoflagellates and microzooplankton species. From 20 to 50 mL aliquots of

samples are settled in separable counting chambers for at least 24 hours before examination under

phase-contrast optics with an inverted microscope following the classic Utermöhl technique (Lund et al.

1958). A single transect across the center of the chamber is counted at 390× magnification for
Nutrient Criteria—Estuarine and Coastal Waters 4-17

flagellates; 150× magnification is used for other organisms. From 25% to 100% of the chamber bottom

is examined, depending on cell concentrations in the sample. Appropriate multipliers are used to convert

all counts to common units of cells/L (Sournia 1978). Organisms are identified to the lowest taxonomic

category possible. Even quite small changes in the physical and chemical parameters and availability of

micronutrients can have a significant effect on the growth constants of algae. A difference in doubling

time of 25% between two fast-growing organisms can lead to one outnumbering the other by 15 to 1 in a

week and quickly lead to alterations in species assemblages (Riley and Chester 1971).

There are numerous algal species in estuarine and open coastal waters that are considered to be harmful

(e.g., see Dortch et al. 1998, Anderson and Garrison 1997, Anderson 2000). This is a rapidly changing

area of marine ecology and experts should be consulted for specific taxonomic identifications.

Macrobenthos, Macroalgae, and Seagrasses and SAV

Macroinfauna are typically sampled with coring devices or bottom grab samplers and wet-sieved through

0.5 μM mesh sieves to separate the animals from very fine sediments. Stacked sieves can be used to

remove larger shell fragments and sand particles. A relaxant (e.g., 0.3% propylene phenoxytol) is

applied prior to addition of formalin. Samples are usually preserved in 10% buffered formalin for several

weeks and then transferred to 60%-70% isopropanol (Diaz and Rosenberg 1995).

Macroalgae are typically sampled by collecting algal material by hand from a known surface area of the

habitat. Various devices may be used (e.g., 0.5 m stainless or plastic hoop).

Both above- and below-ground seagrass and SAV biomass can be collected from a known area of the

bed. Various techniques have been used. An often-used method is to shove metal strips along the

sediment surface in a square meter pattern and anchor the strips at all four corners by pushing a sharp

spike through holes drilled at each end of the strips. Then, the plant material separated to species can be

clear-cut with sharp shears and taken to the laboratory and dried in a heated cabinet at 60°C to constant

dry weight. A sharp spade is required to collect below-ground roots and rhizomes. This material should

be identified and dried to constant weight.
Nutrient Criteria—Estuarine and Coastal Waters 5-1

Developing National/Regional Databases

Sampling Design 

Quality Assurance/Quality Control

Statistical Analyses

CHAPTER 5

Databases, Sampling Design,

and Data Analysis

5.1 INTRODUCTION

Development of national regional numeric nutrient criteria requires that an extensive amount of data

from across the country be evaluated. This information can be an invaluable tool to States and Tribes as

they develop nutrient criteria. Both existing and historical data may provide considerable information

that is specific to the region where criteria are to be set. First the data must be located, then the

suitability of the data (type and quality) ascertained before they can be used for analysis of water quality

parameters. It is also important to determine how the data were collected to make future monitoring

efforts compatible with earlier approaches. Descriptive data that characterize the waterbody are

invaluable.

Data may come from existing sources or can be collected from new sampling programs. Nutrient-related

data for estuaries and coastal waters, collected by various agencies for many different purposes, exist in

numerous databases and have the potential to provide the basis for development of nutrient criteria on a

regional level. This chapter presents an overview of existing databases and a general discussion

concerning the evaluation of such datasets in terms of their use in the nutrient criteria development

process. The list of databases is not all-inclusive—many other data sources exist—but the list provided

is intended to represent the kind of information that is available. This chapter also provides a description

of existing data resources (e.g., U.S. EPA Legacy STORET and ODES) and how these data may be used

to generate preliminary nutrient criteria on regional levels. In addition to discussing the use of existing

data, the chapter discusses new data collection, including consideration for sampling design and the types

of sampling to be considered as part of data collection activities. The chapter ends with a general

discussion of data management, quality assurance, and quality control issues that are integral in the

overall discussion of data storage, accessibility, and utilization. 

5.2 DEVELOPING REGIONAL AND NATIONAL DATABASES FOR ESTUARIES

AND COASTAL WATERS

A database is a collection of information related to a particular subject or purpose. Databases are

arranged so that they divide data into separate electronic repositories in tabular format. Data in tables

can be viewed and edited, and new data can be added. A single datum is stored in only one table but can

be viewed from multiple locations. Updating one view of a datum will update it in all the various

viewable forms. Each table should contain a specific type of information. Data from different tables can

be viewed simultaneously according to the user-defined table relationships. That is, the relationship

among data in different tables can be defined so that more than one table can be queried or reported and

accessed in a single view. Data stored in tables can be located and retrieved using queries. A query
5-2 Nutrient Criteria—Estuarine and Coastal Waters

allows the user to find and retrieve only the data that meet user-specified conditions. Queries also can be

used to update or delete multiple records simultaneously and to perform built-in or custom calculations of

data. Data in tables can be analyzed and printed in specific layouts for reports. 

To facilitate data manipulation and calculations, it is highly recommended that historical and present-day

data be transferred to a relational database. A relational database is a collection of data items organized

as a set of formally described tables from which data can be accessed or reassembled in many different

ways without having to reorganize the database tables. Each table contains one or more data categories

in columns. Each row contains a unique instance of data for the categories defined by the columns. The

organization of data into relational tables is known as the logical view of the database. Relational

databases are powerful tools for data manipulation and initial data reduction. They allow selection of

data by specific and multiple criteria and definition and redefinition of linkages among data components. 

Geographic information systems (GIS) are geo-referenced databases that have a geographic component

(i.e., spatial platform) in the user interface. Spatial platforms associated with a database allow

geographic display of sets of sorted data and make mapping easier. These types of databases with spatial

platforms are becoming more common. The system is based on the premises that “a picture is worth a

thousand words” and that most data can be related to a map or other easily understood graphic. GIS

platforms such as ArcView, ArcInfo, and MapInfo are frequently used to integrate spatial data with

monitoring data for watershed analysis.

The EPA National Nutrient Criteria Program initiated the development of a national database application

that will be used to store and analyze nutrient data. The ultimate use of these data will be to derive

ecoregion- and waterbody-type specific numeric nutrient criteria. Initially, EPA developed a Microsoft

Access application that was populated with STORET Legacy data, U.S. Geological Survey (USGS)

(NAWQA, NASQAN, and Benchmark) data, and other relevant nutrient data from universities,

States/Tribes, and additional data-rich entities. To serve the general public more effectively and

efficiently, EPA also developed and maintains a web-accessible nutrient database application in an

OracleTM environment that allows for easy web accessibility, geo-referencing/GIS compatibility, and data

analysis on a State/Tribal, regional, and national basis. The total amount of existing nutrient data

nationally is large (>20 gigabytes), and it is anticipated that more data will be entered into the system. 

The OracleTM application can easily manage large quantities of data and provides ample room for

expansion as more data are collected. The OracleTM database application is being designed for

compatibility with EPA’s modernized STORET. A key feature of the database design will prevent

duplication of effort for users of STORET and the nutrients database application, especially for data

updating. Considerable efforts are also being made to ensure compatibility with other database systems

(e.g., WQS and RAD) currently being developed by EPA’s Office of Water. The OracleTM application

has been online since the fall of 2000.
Nutrient Criteria—Estuarine and Coastal Waters 5-3

Data Sources

Potential sources of data include water quality monitoring data from Federal, State, Tribal, and local

water quality agencies; university studies; and volunteer monitoring programs. However, the data

sources described in this section do not encompass the full extent of available data sources. The data

available in the nutrient database can be used to identify reference areas to begin development of

potential nutrient criteria. The nutrient data sources for estuaries and coastal waters that will be useful

for developing criteria are discussed below. These data sources contain extensive water quality data,

however, data; collection should not be limited to these sources. Collection of scientifically sound water

quality data from any reliable source is encouraged.

Many of the water quality programs listed here include rivers and streams data or mixed freshwater,

estuarine, and coastal water systems data. The rivers and streams information is included in this

document because it gives relevant data about nutrient loading from fluvial systems, which is important

to estuaries and coastal waters. Generally, in estuaries that have been impaired by nutrients, a database

exists, and in less impaired estuaries, the database is often insufficient for comparisons. Nutrient

loading information from fluvial systems may provide a basis for comparison between systems if they

share important geophysical conditions. Such comparisons would assist in developing trends and

extrapolating where insufficient data exist.

EPA Water Quality Data

EPA has many programs of national scope that focus on collection and analysis of water quality data. 

The following information on several of the databases and national programs may be useful to water

quality managers as they compile data for criteria development.

STOrage and RETrieval System (STORET)

STORET is EPA’s national database for water quality and biological data. EPA’s original STORET

system, called the STORET Legacy Data Center (LDC) and operated continuously since the 1960s, was

historically the largest repository of water quality data in the Nation. This legacy mainframe-based

system was the repository of all data held in EPA’s original STORET system as of the end of 1998. This

Legacy STORET ceased to exist in the year 2000. In its place, EPA is supporting a modernized database,

simply called STORET, designed as a replacement for the original STORET System. While STORET

will serve as the major repository for more current data, the nutrient criteria database application will

offer major improvements in database content and capabilities that will enable more detailed data

analysis. 

Interested parties may view both databases on the World Wide Web. For the nutrient database,

capabilities exist to produce printed reports and download data files. Queries for data via the web will be

designed for use by the general public and will require no special training or software. 

STORET is a compendium of data supplied by Federal, State, and local organizations used to evaluate

environmental conditions in the field. The data in STORET are organized by both geographic location

and data ownership. Every field study site is identified by at least one latitude/longitude and, where
5-4 Nutrient Criteria—Estuarine and Coastal Waters

appropriate, also by State/Province, county, drainage basin, and stream reach. Monitoring activities

recorded include field measurements, habitat assessments, water and sediment samples, and biological

population surveys. Records cover the complete spectrum of physical properties, concentrations of

substances, and abundance and distribution of species observed during biological monitoring. STORET

is designed for maximum compatibility with commercial software, including GISs such as the ESRI

ArcView package, and statistical packages such as PC SAS. STORET downloaded files import easily

into all standard spreadsheet packages. Further information about STORET may be obtained by emailing STORET@epa.gov, or telephoning toll-free at 1-800-424-9067.

Environmental Monitoring and Assessment Program (EMAP)

EMAP is an EPA research program designed to develop the tools necessary to monitor and assess the

status and trends of national ecological resources (see EMAP Research Strategy on the EMAP website: 

www.epa.gov/emap). EMAP's goal is to develop the scientific understanding for translating

environmental monitoring data from multiple spatial and temporal scales into assessments of ecological

condition and forecasts of future risks to the sustainability of the Nation’s natural resources. EMAP's

research supports the National Environmental Monitoring Initiative of the Committee on Environment

and Natural Resources (CENR). Data from EMAP can be downloaded directly from the EMAP website. 

The EMAP Data Directory contains information on available datasets, including data and metadata

(language that describes the nature and content of data). The status of the Data Directory as well as

composite data and metadata files also are available on the EMAP website. EMAP–estuaries data is one

of several areas addressed by the program. Most of the estuaries data were collected during a summer

index period.

Ecological Data Application System (EDAS)

EDAS is EPA’s program-specific counterpart to STORET. EDAS was developed by EPA’s Office of

Water to manipulate data obtained from biological monitoring and assessment and to assist States/Tribes

in developing biocriteria. It contains built-in data reduction and recalculation queries that are used in

biological assessment. The EDAS database is designed to enable the user to easily manage, aggregate,

integrate, and analyze data to make informed decisions regarding the condition of a water resource.

Biological assessment and monitoring programs require aggregation of raw biological data (lists and

enumeration of taxa in a sample) into informative indicators. EDAS is designed to facilitate data

analysis, particularly the calculation of biological metrics and indexes. Predesigned queries that

calculate a wide selection of biological metrics are included with EDAS. Future versions of EDAS will

include the capability to upload data to, and download data from, the distributed version of modernized

STORET. EDAS is not a final data warehouse, but it is a program or project-specific customized data

application for manipulating and processing data to meet user requirements. The EDAS application is

currently under development; more information will be available through the EPA website.

Ocean Data Evaluation System (ODES)

ODES is used for storing and analyzing water quality and biological data from marine, estuarine, and

some freshwater environments. The system supports Federal, State, and local decisionmakers associated

with marine monitoring programs and managers and analysts who must meet regulatory objectives
Nutrient Criteria—Estuarine and Coastal Waters 5-5

through the evaluation of marine monitoring information. ODES contains data from the National Estuary

Program, the Great Lakes National Program Office, the Ocean Disposal Program, the 301(h) Sewage

Discharge Program, the National Pollutant Discharge Elimination System Program, and the 403(c)

Program. Records pertain to water quality, fish abundance, bioaccumulation, benthic infauna, fish

histopathology, bioassay, and sediment physical/chemical characteristics. Users can examine both

spatial and temporal relationships among variables. A quality assurance report describing analytical

methods and procedures for each dataset is stored with each dataset.

Chesapeake Bay Program (CBP)

CBP, a cooperative effort between the Federal Government, the States, the District of Columbia, and

local governments in the Chesapeake Bay watershed, provides funds to the States of Maryland and

Virginia for the routine monitoring of 19 directly measured water quality parameters at 49 stations within

the bay watershed. The Water Quality Monitoring Program began in June 1984 with stations sampled

once each month during the colder late fall and winter months and twice each month in the warmer

months. A refinement in 1995 reduced the number of monitoring cruises to 14 per year. Data are

available on the internet at www.chesapeakebay.net/data/. 

National Estuarine Programs (NEPs)

Many NEPs have nutrient and related data that could be used for characterization purposes. Presently,

there is no national repository of NEP data, but, in the development of regional nutrient criteria, the

NEPs may serve as an excellent source for information. Some of these programs have electronic

databases and some hard copy data that could be acquired. EPA is attempting to acquire the available

NEP data and eventually enter them into the National Nutrient Criteria Program database. A list of NEP

estuarine systems can be found online at www.epa.gov/nep.

National Oceanographic and Atmospheric Administration (NOAA) 

Water Quality Data in the National Oceanographic Data Center (NODC)

NODC is one of three national environmental data centers operated by NOAA and serves as a national

repository and dissemination facility for global oceanographic data. Its primary mission is to ensure that

global oceanographic data collected at great cost are maintained in a permanent archive easily accessible

to the world science community and to other users. NODC holds physical, chemical, and biological

oceanographic data collected by U.S. Federal agencies, including the Department of Defense (primarily

the U.S. Navy); State and local government agencies; universities and research institutions; and private

industry. NODC does not conduct any data collection programs of its own; it serves solely as a

repository and dissemination facility for data collected by others (see website at www.nodc.noaa.gov). 

NODC provides data management support for major ocean science projects such as Tropical

Ocean–Global Atmosphere (TOGA), World Ocean Circulation Experiment (WOCE), and Joint Global

Ocean Flux Study (JGOFS). NODC's global holdings of physical, chemical, and biological

oceanographic data include substantial amounts of data from coastal ocean areas. For example, the

NODC Oceanographic Profile Database holds primarily coastal data (www.nodc.noaa.gov/cgibin/JOPI/jopi).
5-6 Nutrient Criteria—Estuarine and Coastal Waters

National Estuarine Research Reserve System (NERR)

The NERR Systemwide Monitoring Program was designed to identify and track short-term variability

and long-term changes in the integrity and biodiversity of representative estuarine ecosystems and coastal

watersheds for the purposes of contributing to effective national, regional, and site-specific coastal zone

management. The program has two major goals: (1) to support State-specific nonpoint source pollution

control programs by establishing local networks of continuous water quality monitoring stations in

representative protected estuarine ecosystems; and (2) to develop a nationwide database on baseline

environmental conditions in the NERR system of estuaries. Water quality data collected from phase 1 of

the NERR Systemwide Monitoring Program provides data necessary for site and intersite baseline

studies, trend analysis, and impact assessment. Data are available for each of the participating NERR

systems at http://inlet.geol.sc.edu/cbmoweb/30_minute_data.html.

Rivers and Streams Water Quality Data

Rivers and streams water quality data are potentially useful for estuaries and coastal waters. Because

much of the nutrient load to estuaries comes from rivers and streams, it is critical to define nutrient

concentrations landward of tidal influence and to calculate fluvial-based nutrient loads to estuaries and

potentially to coastal waters. EPA STORET, which was discussed in detail previously, includes data

from rivers and streams from across the Nation. Another comprehensive Federal source of river and

stream water quality data is USGS. USGS maintains databases on water quantity and quality for

waterbodies across the Nation. Many of the data for rivers and streams are available through the

National Water Information System (NWIS). The most convenient method of accessing the local

databases is through the USGS State representative. Every State office can be reached through the USGS

home page on the Internet at URL http://www.usgs.gov/wrd002.html. The USGS data from several

national water quality programs covering large regions offer highly controlled and consistently collected

data that may be particularly useful for nutrient criteria analysis. Two programs, the Hydrologic

Benchmark Network (HBN) and the National Stream Quality Accounting Network (NASQAN), include

routine monitoring of rivers and streams during the past 30 years. The USGS National Water Quality

Assessment (NAWQA) Program is building a third national database of stream quality data collected and

analyzed for more than 50 river basins and aquifer systems across the Nation. More information and data

from each of these studies can be found on the USGS website (www.usgs.gov). For additional data

sources, the Rivers and Streams Nutrient Criteria document presents an extensive list of related

freshwater nutrient-related information (Nutrient Criteria Technical Guidance Manual—Rivers and

Streams, 2000, EPA-822-B-00-002). 

USGS San Francisco Bay Program

Since 1968, USGS has sustained a research program to understand how coastal ecosystems function and

how those functions are altered by human disturbances. One component of this program is directed to

following and understanding changes in the water quality of San Francisco Bay. The program includes

regular measurements of water quality along a 145-kilometer transect spanning the length of the entire

estuarine system, from the South Bay to the Sacramento River. The program studies many different

aspects of San Francisco Bay, such as changing land use, hydrology, water currents, nutrients, toxic
Nutrient Criteria—Estuarine and Coastal Waters 5-7

contaminants, geological structure, and biological communities. The results of water quality

measurements, and eventually the full dataset, can be accessed at http://sfbay.wr.usgs.gov/access/index/

wqdata.html.

State/Tribal Monitoring Programs

Most States monitor some estuaries and coastal waters within their borders for algal and nutrient

variables. Data collected by State/Tribal water quality monitoring programs can be used for nutrient

criteria development. These data should be available from the agencies responsible for monitoring.

Sanitation Districts

Massachusetts Water Resources Authority (MWRA)

MWRA has conducted a comprehensive monitoring program in Boston Harbor, Massachusetts Bay, and

Cape Cod Bay from 1992 to the present. The program was established to understand baseline conditions

and monitor the effects of effluent discharges into Boston Harbor and Massachusetts Bay. This

multifaceted monitoring program focuses on water quality, benthic ecosystem health, effluent

characterization, and public health issues related to metal, organic, and microbiological contaminants. 

All data are stored in an OracleTM Relational Database Management System to support the monitoring

program.

New York City–Department of Environmental Protection (NYCDEP)

NYCDEP has conducted extensive monitoring to evaluate viable treatment options for sewage effluents

to mitigate conditions that promote eutrophication. The city’s monitoring programs have included point

source, water column, sediment, hydrodynamic and atmospheric studies. All data are stored in the

NYCDEP databases and have been used in the development and application of a System-Wide

Eutrophication Model (SWEM) to enhance the city's ability to evaluate the effectiveness of various

treatment options in mitigating conditions that promote eutrophication.

Southern California 

The major southern California dischargers of treated sewage effluents into marine waters have conducted

applied research and monitoring programs for more than 30 years. These dischargers include the cities of

San Diego and Los Angeles and the counties of Ventura/Oxnard, Los Angeles, Orange, and San Diego. 

The programs are designed to monitor the concentrations and mass emission rates of effluent materials in

the treated effluent; the transport and fate of these materials in the receiving waters; the exposure of the

contaminants to organisms in the receiving waters; and the effects of that exposure to individuals,

populations, and communities of subtidal, intertidal, and water column organisms. Some of this

monitoring is performed to comply with NPDES monitoring efforts, and other monitoring addresses

specific issues of interest to the districts. The data are retained on a number of local databases, but they

are also maintained on EPA’s ODES. In addition to the localized databases managed by the sanitation

districts, a research organization (Southern California Coastal Water Research Project) has performed

parallel and specialized monitoring and applied research on the effects of treated sewage effluents in this

region since the early 1970s. Their data are managed onsite and are provided to national data inventories

(e.g., ODES, NODC, STORET).
5-8 Nutrient Criteria—Estuarine and Coastal Waters

California and Oregon 

Similar monitoring requirements are established for other locations in California and Oregon. The

districts that have effluents that are discharged into rivers, streams, or oceans are required, through the

NPDES permits, to monitor their treated effluents and the receiving waters. The data are retained

locally, but must also be filed with ODES. The State and Regional Water Resources Control Boards of

California and the Departments of Ecology in Washington and Oregon administer these permits.

Puget Sound, Washington 

The cities and counties on the Puget Sound watershed monitor the treated sewage effluents and the

receiving waters in compliance with State water quality parameters. These data are provided to the State

in electronic format and are retained on database systems administered by the Department of Ecology

that are available to ODES. The data in the receiving water environments are collected in methods that

are historically similar to work that has been performed in Puget Sound since the middle 1960s. 

Academic and Literature Sources

Many research studies conducted by academic institutions may provide data useful for developing

nutrient criteria. Academic research tends to be site specific and span a limited number of years,

although data for some systems may span 20 years or more. Academic research data should be available

from researchers. However, the scientific literature is likely to be a major source of estuarine and coastal

waters data.

Volunteer Monitoring Programs

Many States have volunteer water quality monitoring programs. Some programs are State sponsored,

while others are independently organized. Citizens in many areas donate their time, money, or

experience to aid State, Tribal, and local governments in collecting water quality data. Volunteers

analyze water samples for dissolved oxygen, nutrients, pH, temperature, and a host of other water

constituents; evaluate the health of stream habitats and aquatic biological communities; note shore zone

conditions and land uses that may affect water quality; catalogue and collect beach debris; and restore

degraded habitats. 

State and local agencies may use volunteer data to screen for water quality problems, establish trends in

waters that would otherwise be unmonitored, and make planning decisions. Volunteers benefit from

learning more about their local water resources and identifying what conditions or activities might

contribute to pollution problems. As a result, volunteers frequently work with clubs, environmental

groups, and State/Tribal or local governments to gather information and address problem areas. As with

any other data source, whether student, State, Federal, academic, or volunteer based, documented quality

assurance procedures are an important consideration.

EPA supports volunteer monitoring and local involvement in protecting our water resources. EPA

support takes many forms, including sponsoring national and regional conferences to encourage

information exchange among volunteer groups, government agencies, businesses, and educators;

publishing sampling methods manuals for volunteers; and providing technical assistance (primarily on
Nutrient Criteria—Estuarine and Coastal Waters 5-9

quality control and laboratory methods) and regional coordination through the 10 EPA Regional Offices. 

EPA also produces a Nationwide Directory of Volunteer Monitoring Programs, which is available online

at http://yosemite.epa.gov/water/volmon.nsf. This directory lists volunteer organizations around the

country engaged in monitoring rivers, lakes, estuaries, beaches, wetlands, and groundwater, as well as

surrounding lands. EPA volunteer monitoring activities are coordinated in part through a website that

lists many resources at http://www.epa.gov/owow/monitoring/volunteer.

Quality of Historical Data

The value of older historical datasets is a recurrent problem because data quality is often unknown. 

Knowledge of data quality is also problematic for long-term data repositories such as STORET and longterm State databases, where objectives, methods, and investigators may have changed many times over

the years. The most reliable data tend to be those collected by a single agency, using the same protocol,

for a limited number of years. Supporting documentation should be examined to determine the

consistency of sampling and analysis protocols. Investigators must determine the acceptability of data

contained in large, heterogeneous data repositories. Considerations and requirements for acceptance of

these data are described below.

Location Data

STORET and USGS data are geo-referenced with latitude, longitude, and up to Reach File 3 (RF1 &

RF3) codes. Geo-reference data can be used to select specific locations or specific USGS hydrologic

units. In addition, STORET often contains a site description. Knowledge of the rationale and methods

of site selection from the original investigators may supply valuable information. Metadata of this type,

when known, are frequently stored within large long-term databases.

Variables and Analytical Methods

Thousands of variables are recorded in database records. Each separate analytical method yields a

unique variable. For example, five ways of measuring total nitrogen (TN) results in five unique

variables. We do not recommend mixing analytical methods in sample analyses because methods differ

in accuracy, precision, and detection limits. Sample analyses should concentrate on a single analytical

method for each parameter of interest. Selection of a particular “best” method may result in too few

observations, in which case it may be more fruitful to select the most frequently used analytical method

in the database. Data may have been recorded using analytical methods under separate synonymous

names, or analytical methods incorrectly entered when data were first added to the database. Review of

recorded data and analytical methods recorded by knowledgeable personnel is necessary to correct these

problems. 

Laboratory Quality Control

Laboratory quality control of data (blanks, spikes, replicates, known standards, etc.) where available

should be reported. Such information may have been infrequently reported in larger data repositories and

needs to be identified and coded. Records of general laboratory quality control protocols and specific

quality control procedures associated with specific datasets are valuable in evaluating data quality. 

However, premature elimination of lower quality data can be counterproductive because the increase in
5-10 Nutrient Criteria—Estuarine and Coastal Waters

variance caused by analytical laboratory error may be negligible compared with natural variability or

sampling error, especially for nutrients and related water quality parameters. However, data of uncertain

quality should not be accepted unless no other data are available.

Data Collecting Agencies

Selecting data from particular agencies with known, consistent sampling and analytical methods will

reduce variability caused by unknown quality problems. Requesting data review for quality assurance

from the collecting agency will reduce uncertainty about data quality.

Time Period

Long-term records are critically important for establishing trends. Determining if trends exist in the time

series database is also important for characterizing reference conditions for nutrient criteria. Length of

time series data needed for analyzing nutrient data trends is discussed in the Sampling Design section

(Section 5.3).

Index Period

The index period for estimating average concentrations can be established if nutrient and water quality

variables were measured through seasonal cycles. The index period may be the entire year or the

summer season. The best index period is determined by considering water quality characteristics for the

region, the quality and quantity of data available, and estimates of temporal variability (if available). 

Additional information and considerations for establishing an index period are discussed in Section 5.3.

Representativeness

Data may have been collected for specific purposes. Data collected for toxicity analyses, effluent limit

determinations, or other pollution problems may not be useful for developing nutrient criteria. 

Furthermore, data collected for specific purposes may not be representative of the spatial scale of

interest. The investigator must determine if the spatial scale for the data included in the database is

representative of the area to be characterized. If a sufficient amount of data for the appropriate scale

cannot be found, then new surveys will be necessary (see Section 5.3).

Gathering New Data

New data should be gathered following the sampling design protocols discussed below. New data

collection activities for developing nutrient criteria should focus on filling in gaps where data are

particularly needed for high-risk systems. Data gathered under new monitoring programs should be

imported into databases or spreadsheets and merged with the existing nutrient database for criteria

development. 

5.3 SAMPLING DESIGN

This section discusses issues surrounding sampling nutrients, response variables, and related

environmental variables in estuaries and coastal waters. Where appropriate data are unavailable or

insufficient to derive numerical nutrient criteria, efforts must be made to collect new data to fill those
Nutrient Criteria—Estuarine and Coastal Waters 5-11

gaps. New sampling programs should be scientifically based and statistically rigorous while maximizing

available management resources. Such programs are used to better define nutrient and algal relationships

within an ecosystem framework. At the broadest level, sampling efforts should detect or contribute to the

following objectives:

C Identify the reference condition, that is, existing, most natural, least culturally impacted locations

and their relative enrichment status

C Identify whether nutrient concentrations or loads are increasing, decreasing, or remaining the same

C Characterize seasonal patterns in nutrient levels and their relationship to primary productivity

C Help assess the assimilative capacity of the system, that is, contribute to the determination of how

much nutrient loading can be assimilated without causing unacceptable changes in water quality or

algal biomass and composition. (Note: In estuaries and coastal waters, this objective will likely

require application of a computer-based hydrodynamic and nutrient-coupled water quality model. 

The intention here is to recognize particularly susceptible waters, not to lower expectations relative

to historical antecedents and the reference condition.)

Some sampling programs may be poorly and inconsistently funded or are improperly designed and

carried out, making it difficult to collect a sufficient number of samples over time and space to identify

changes in water quality or to estimate average conditions with statistical rigor. This section provides a

procedural approach for assessing water quality condition and identifying impairment by nutrients and

algae in estuaries and coastal waters. The approaches described below present sampling designs that

allow one to obtain a significant amount of information while attempting to minimize overall effort (and

cost). Probabilistic and stratified random sampling begin with large-scale random monitoring designs

that are reduced as nutrient and algal conditions are characterized. The tiered approach to sampling

begins with coarse screening and proceeds to more detailed protocols as impaired and high-risk systems

are identified and targeted for further investigation. 

Sampling Protocol

Success of nutrient criteria development requires that consideration be given to sampling design. 

Initially, the relationship between critical response variables and nutrient concentrations, or in some

cases, nutrient loads, needs to be established. Next, reference sites should be sampled, if feasible, in an

attempt to establish reference conditions within classes of systems or subsystems. Classification should

be linked to the reference condition activity. Nutrient concentrations/load and algal biomass

relationships should help define the ecological state that can be attained if impaired systems are restored. 

As discussed in the following sections, this is not a straightforward exercise; it is very difficult to predict

water transport/mixing in estuaries and coastal waters. The physics of these waters plays a major role in

determining the observed patterns in nutrient and chlorophyll a concentrations, turbidity, and bottom

water dissolved oxygen deficiency as well as transport. Variability in time and space further complicates

empirical analysis as pointed out in Chapter 2.
5-12 Nutrient Criteria—Estuarine and Coastal Waters

Nutrient concentrations, chlorophyll a production, and system respiration represented as biological

oxygen demand (BOD) are biochemical processes. How these processes are expressed in terms of

dissolved oxygen, especially in bottom waters, has a great deal to do with the variability in freshwater

flows, density stratification, advection, and mixing. The forcing factors include wind setups, changes in

barometric pressure gradients, freshwater gravitational circulation, and the added complexity of bottom

bathymetry. Interactions of these factors may create “flow jets,” flow reversals, three-layered circulation,

and other physical complexities that suggest any monitoring scheme planned for estuaries and coastal

waters would be advised to have a physical oceanographer as part of the team. 

Sampling Technique

A reasonable and representative method is to profile the general physical character of the site by a CTD

hydrocast. Water samples then may be collected from the surface 1 meter, mid-depth, and bottom 1

meter of the water column. Sample station activities should be coordinated as much as possible with the

same tidal and current phase each time data are collected. If turbidity is measured by Secchi depth, the

disc should be lowered from the shaded side of the vessel and depth determined from as close to the

water level as practicable. Secchi depth measurements should be made only during periods of full

daylight. The data for each station and sampling event should be recorded for each depth interval. This

permits assessment for surface as well as bottom conditions. Where satisfactory, the results for each

sampling episode can be combined into a mean or median measure representing all depths at that site. 

Temporal and spatial medians of the sites then can be determined to establish the representative values

for that reference site.

During sampling visits, the candidate reference stations also should be examined to confirm whether they

actually meet the reference site requirements. This may include looking for nearby discharges into the

waters or tributaries and a quick survey of the shoreline to determine if new modifications may have

changed the site. If an area appears to have been significantly impacted, measurements should be made

for nutrient concentrations and biological response variables such as chlorophyll concentration and fish,

macroinvertebrate, macrophyte, and planktonic community variables. Sites that do not meet the physical

reference requirements should be excluded from the reference dataset. However, a high nutrient

concentration present in an otherwise minimally developed area is not justification alone for exclusion. 

This may be part of the natural background level to be identified by the reference condition process.

Initial Considerations

Variability is inherent in sampling, which means that accuracy (how well the measure reflects actual

conditions) and precision (how consistent the measurement is) must be assessed. Precision in ecological

samples and measurements is more easily characterized than accuracy. Replicate samples from an

experimental unit provide the basis for precision analysis. Standard statistical textbooks focus on

precision (i.e., various ways to assess the nature of variability) and especially inferences regarding null

hypotheses. In analytical chemical analyses, accuracy can be assessed by including samples of known

purity and/or amount. Accuracy often refers to systematic errors in a method, whereas precision refers

more or less to random errors. Outliers can be detected with statistical methods, but the so-called outlier

may actually prove to be more accurate than the remainder of the data.
Nutrient Criteria—Estuarine and Coastal Waters 5-13

Some key questions to consider are the following: Does the method reflect more on how the assessor or

analyst uses the method if the technician is imprecise, and Does the method actually do what it is

purported to do? For example, is the arsenic concentration high enough that it interferes importantly

with the phosphate analysis? Does the Loran system accurately place one on station, especially when

electronic interferences are highly probable? Does the temperature setting on the dissolved organic

carbon analyzer allow for accurate measurement or a biased one, or does the chlorophyll a method

accurately correct for interferences from other pigments? For these examples, the precision may be very

high (low variance) but the measurement may be in substantial error so the results are inaccurate. The

following discussion expands on these ideas and provides additional information on statistical concepts

and procedures relative to sample design. It is worth remembering that environmental data may not

conform to the assumptions of normality required for statistical inferences; adjustments often can satisfy

parametric assumption, but when they do not, the analysts must resort to distribution-free methods.

Specifying the Population and Sample Unit

Sampling is statistically expressed as a sample from a population of objects. Finite populations may be

sampled with corresponding natural sample units, but often the sample unit (say, an estuary) is too large

to measure in its entirety, and it must be characterized with one or more second-stage samples of the

sampling gear (bottles, benthic grabs, quadrats, etc.). Each sample unit is assumed to be independent of

other sample units. The objective of sampling is to best characterize individual sample units in order to

estimate some attributes (e.g., nutrient concentrations or dissolved oxygen) and their statistical

parameters (e.g., mean, median, variance and percentiles) of a population of sample units. The objective

of the analysis is to be able to say something (estimate) about the population. It is critical to distinguish

between making an inference about a population of many estuaries (e.g., “lagoonal estuaries around

northern Gulf of Mexico are shallow and mesotrophic”) versus an inference about a single estuary or

coastal water (e.g., “estuary ABC has fewer fish species than unimpaired reference estuaries or salinity

zones within estuaries”). These two kinds of inferences require different sampling designs: the first

requires independent observations of many waterbodies and does not require repeated observations

within sample units (pseudoreplication; Hurlbert 1984), while the second often does require repeated

observations within a waterbody. Examples of sample units include:

C A point in an estuary or coastal water (may be characterized by single or multiple sample device

deployments). The population then would be all points in the waterbody, an infinite population.

C A constant area (e.g., square meter, hectare). The population could be all square meters of a coastal

water surface area in a State or region.

C An estuary or a definable subbasin or salinity zone of an estuary as a single unit. Because salinity

often specifies population distributions in estuaries, these zones most often are discrete

environments, at least in the short term, and this is likely to be the most common sample unit. The

population would be all salinity zones in a State or region, a finite population.
5-14 Nutrient Criteria—Estuarine and Coastal Waters

Specifying the Reporting Unit

It is also necessary to specify the units for which results will be reported. Usually, this is the population

(e.g., all estuaries), but it also can be subpopulations (e.g., estuaries within a given nutrient ecoregion)

and even individual locations (e.g., estuaries or coastal waters of special interest). To help develop the

sampling plan, it is useful to create hypothetical statements of results in the way that they will be

reported, for example:

C Status of a place: “The estuary ABC is degraded.”

C Status of a region: “An estimated 20% of the estuary area in State XYZ has an elevated trophic

state, above reference expectations”; “Approximately 20% of estuaries in State XYZ have an

elevated trophic state.”

C Trends at a place: “Nutrient concentrations in estuary ABC have decreased by 20% since 1980.”

C Trends of a region: “Average estuary trophic state in State XYZ has increased by 20% since

1980”; “Average trophic state index values in 20% of estuaries of State XYZ have increased by

15% or more since 1980.”

C Relationships among variables: “A 50% increase of N loading above natural background is

associated with decline in taxa richness of benthic macroinvertebrates, below reference

expectations.”; “Coastal waters receiving runoff from large nonpoint sources have 50% greater

probability of elevated trophic state above reference conditions than coastal waters not receiving

such runoff.”

Sources of Variability

Variability of measurements has many possible sources, and the intent of many sampling designs is to

minimize the variability due to uncontrolled or random effects, and conversely to be able to characterize

the variability caused by experimental or class effects. For example, estuaries may be classified by soil

phosporus content of their surrounding watersheds so that estuaries within a class are likely to have

similar water column concentrations in current or historical reference areas. The population of estuaries

is stratified so that observations (sample units) from the same stratum will be more similar to each other

than to sample units in other strata.

Environmental measures vary across different scales of space and time, and sampling design must

consider the scales of variation. In coastal waters, measurements of some variables such as total nitrogen

or chlorophyll concentrations are taken at single points in space and time (center of the deep depression,

20 m depth, 10 a.m. on 2 July). If the same measurement is taken at a different place (littoral zone, 1 m),

or coastal waters, or time (30 January), the measured value may be different. A third component of

variability is the ability to accurately measure the quantity of interest, which can be affected by sampling

gear, instrumentation, errors in proper adherence to field and laboratory protocols, and the choice of

methods used in making determinations.
Nutrient Criteria—Estuarine and Coastal Waters 5-15

The basic rule of efficient sampling and measurement is to sample so as to minimize measurement errors,

to maximize the components of variability that have influence on the central questions and reporting

units, and to control other sources of variability that are not of interest, that is, to minimize their effects

on the observations. In the example of chlorophyll concentrations, variability could be reduced by

sampling each of several coastal waters in the deepest part, with multiple depth samples or a vertically

integrated pump sample taken in early spring before stratification appears. Many coastal waters are

sampled to examine and characterize the variability due to different coastal waters (the sampling unit). 

Each coastal water is sampled in the same way, in the same place, and in the same timeframe to minimize

variability due to location, depth, and season, which are not of interest in this particular study.

In the above example, chlorophyll concentrations vary with location within a coastal water, among

coastal waters, and time of sampling (day, season, year). If the spatial and temporal components of

variability within coastal waters are large (e.g., measurements of chlorophyll concentrations typically

vary more between spring and fall samples within a coastal water than they do among coastal waters),

then it may be best to use an index period. For this reason, coastal water chlorophyll concentrations

often are estimated as a growing season average, estimated from several determinations (e.g., monthly)

during the growing season.

In statistical terminology, there is a distinction between sampling error and measurement error that has

little to do with actual errors in measurement. Sampling error is the error attributable to selecting a

certain sample unit (e.g., a coastal water or a location within a coastal water) that may not be

representative of the population of sample units. Statistical measurement error is the ability of the

investigator to accurately characterize the sampling unit. Thus, measurement error includes components

of natural spatial and temporal variability within the sample unit as well as actual errors of omission or

commission by the investigator. Measurement error is minimized with methodological standardization:

selection of cost-effective, low-variability sampling methods; proper training of personnel; and quality

assurance procedures to minimize methodological errors. In analytical laboratory procedures,

measurement error is estimated by replicate determinations on some subset of samples (but not

necessarily all). Similarly, in field investigations, some subset of sample units should be measured more

than once to estimate measurement error.

Analysis of variance (ANOVA) can be used to estimate measurement error. All multiple observations of

a variable are used (from all coastal waters with multiple observations), and coastal waters are the

primary effect variable. The root means square error (RMSE) of the ANOVA is the estimated variance

of repeated observations within coastal waters. Note that a hypothesis test (F-test) is not of interest in

this application, only the RMSE of the analysis.

Natural variability that is not of interest for the questions being asked, but which may affect the ability to

address them, should be estimated with the RMSE method above. If the variance estimated from RMSE

is unacceptably large (i.e., as large or larger than variance expected among sample units), then it is often

necessary to alter the sampling protocol, usually by increasing sampling effort in some way to further

reduce the measurement error. Measurement error can be reduced by multiple observations at each
5-16 Nutrient Criteria—Estuarine and Coastal Waters

sample unit, for example, multiple ponar casts at each sampling event; multiple observations in time

during a growing season or index period; depth-integrated samples; or spatially integrated samples.

A less costly alternative to multiple measures in space is spatially composite determinations. In nutrient

or chlorophyll determinations, a water column pumped sample, where the pump hose is lowered through

the water column, is an example of a spatially composite determination. Spatial integration of an

observation and compositing the material into a single sample is almost always more cost-effective than

retaining separate, multiple observations. This is especially so for relatively costly laboratory analyses

such as organic contaminants and benthic macroinvertebrates, but the price of this economy is loss of

information about the water column or about distribution over an area.

Statistical power is the ability of a given hypothesis test to detect an effect that actually exists and must

be considered when designing a sampling program (e.g., Peterman 1990, Fairweather 1991). The power

of a test (1-b) is defined as the probability of correctly rejecting the null hypothesis when the null

hypothesis is false (i.e., the probability of correctly finding a difference [impairment] when one exists).

For a fixed confidence level (e.g., 90%), power can be increased by increasing the sample size or the

number of replicates, except in cases where the variance is proportional to the mean. To evaluate power

and determine sampling effort, an ecologically meaningful amount of change in a variable must be set.

Optimizing sampling design requires consideration of tradeoffs among the measures used, the effect size

that is considered meaningful, desired power, desired confidence, and resources available for the

sampling program. Every study requires some level of repeated measurement of sampling units to

estimate precision and measurement error. Repeated measurement at 10% of sites is common among

many monitoring programs.

Alternative Sampling Designs

Sampling design is the selection of a part of a population to observe the attributes of interest to estimate

the values of those attributes for the whole population. Classical sampling design makes assumptions

about the variables of interest; in particular, it assumes that the values are fixed (but unknown) for each

member of the population until that member is observed (Thompson 1992). This assumption is perfectly

reasonable for some variables, say, length, weight, and sex of members of an animal population, but it

seems less reasonable for more dynamic variables such as nutrient concentrations, loadings, or

chlorophyll concentrations of estuaries. Designs that assume that the observed variables are themselves

random variables are model-based designs, where prior knowledge or assumptions are used to select

sample units.

Probability-Based Designs (Random Sampling)

The most basic probability-based design is simple random sampling, where all possible sample units in

the population have the same probability of being selected; that is, all possible combinations of n sample

units have equal probability of selection from among the N units in the population. If the population N is

finite and not excessively large, a list can be made of the N units, and a sample of n units is randomly
Nutrient Criteria—Estuarine and Coastal Waters 5-17

selected from the list. This is termed list frame sampling. If the population is very large or infinite (such

as locations in an estuary), one can select a set of n random (x,y) coordinates for the sample.

All sample combinations are equally likely in simple random sampling, thus there is no assurance that the

sample actually selected will be representative of the population. Other unbiased sampling designs that

attempt to acquire a more representative sample include stratified, systematic, multistage, and adaptive

designs. In stratified sampling, the population is subdivided or partitioned into strata, and each stratum is

sampled separately. Partitioning is typically done so as to make each stratum more homogeneous than

the overall population; for example, estuaries could be stratified on ecoregion or coastal waters by

dominant current structure. Systematic sampling is the systematic selection of every kth unit of the

population from one or more randomly selected starting units, and it ensures that samples are not

clumped in one region of the sample space. Multistage sampling requires selection of a sample of

primary units, such as fields or hydrologic units, and then selection of secondary sample units, such as

plots or estuaries within each primary unit in the first-stage sample.

Estimation of statistical parameters requires weighting of the data with inclusion probabilities (the

probability that a given unit of the population will be in the sample) specified in the sampling design. In

simple random sampling, inclusion probabilities are by definition equal, and no corrections are

necessary. Stratified sampling requires weighting by the inclusion probabilities of each stratum. 

Unbiased estimators have been developed for specific sampling designs and can be found in sampling

textbooks, such as Thompson (1992).

Model or Goal-Based Designs

Use of probability-based sampling designs may miss relationships among variables (models), especially

if there is a regression-type relationship between an explanatory and a response variable. As an example,

elucidation of estuary response to N loading with the Vollenweider-type model; that is, chlorophyll a

concentration regressed against a depth-normalized N concentration (Vollenweider 1968) requires a

range of trophic states from ultra-oligotrophic to hypereutrophic. A simple random sample of estuaries is

not likely to capture the entire range (i.e., there would be a large cluster of "mesotrophic" estuaries with

few at high or low ends of the trophic scale), and the random sample therefore may be biased with

respect to the model. 

In model-based designs, sites are selected based on prior knowledge of auxiliary variables, such as

estimated phosphorus loading, estuary depth, and elevation. Often, these designs preclude an unbiased

estimate of the population response variable (e.g., trophic state), unless the model can be demonstrated to

be robust and predictive, in which case the population value is predicted from the model and from prior

knowledge of the auxiliary (predictive) variables. Selection of unimpacted reference sites is an example

of samples for a model (index development; response of index variables to measures of anthropogenic

influence) that cannot later be used for unbiased estimation of the biological status of estuaries. Ideally, it

may be possible to specify a design that allows unbiased estimation of both population and model. 

Statisticians should be consulted in developing the sample design for a nutrient criteria program.
5-18 Nutrient Criteria—Estuarine and Coastal Waters

Sampling and Analytical Designs for More Complex Ecological Questions

Complex ecological questions may not be required to develop numerical nutrient criteria. However, the

manager may require a biological and an ecological assessment of resources at risk to establish

management goals—that is, focus on biological resources of high social and economic value (e.g., the

Chesapeake Bay Program includes biological variables as part of the goal setting process). Questions on

how to sample different levels of biological organization (e.g., populations, communities, and

ecosystems), indicators of stress, diversity and similarity measures, and biotic indices may become

important. Criteria that go beyond the core variables will likely address one or more of these ecosystem

or community elements. Publications are available that provide conceptual and statistical guidance for

monitoring biological/ecological systems, including multivariate analytical approaches (e.g., Spellerberg

1991, Luepke 1979, Digby 1987, Clark and Warwick 1994a,b, Ott 1995, Ludwig and Reynolds 1988,

Eckblad 1991).

Monitoring Programs

The purpose of monitoring is to obtain data that can be used not only to determine reference condition,

but to help classify estuaries and coastal waters, or portions thereof, into groups (see Chapter 3). 

Classification should aid in the determination of reference sites or stations that are representative and

have the lowest possible variability.

In some cases, a problem may exist where monitoring data indicate that the system has been greatly

impaired from nutrient enrichment over the period of record. This is analogous to the so-called “cornbelt” problem in some lakes in the upper mid-West of the United States (see Lakes and Reservoirs

Nutrient Guidance Document). This problem suggests that a meaningful reference condition may no

longer exist. Or, the system has been greatly disturbed and it is not clear to what extent the impairments

are due to nutrient enrichment. In this case, both historical information and diagnostic sampling may be

required to clarify the reference condition and subsequent nutrient criteria.

Where data are insufficient, several approaches can be tried, for example, running a mathematical

nutrient model “backwards”; use of biostratigraphic approaches, including changes in algal dominance

and composition, loss of submerged aquatic vegetation (SAV), and pyritization of iron in sediments to

detect earlier anoxia (Brush 1984, 1986, 1992, Cooper 1995) or reference to old written accounts (e.g.,

newspapers, diaries). For example, there are accounts of water clarity in the mouth of the Patuxent River

estuary, Chesapeake Bay, in the late 1930's where engineers sitting in a “Beebe-like Bathysphere” on the

estuary bottom could see horizontally approximately 20 to 30 feet. The methods discussed here are often

qualitative to semiquantitative, but such information can be useful, especially if marked nutrient

increases have evidently occurred over historical conditions but ambient data are insufficient. Older

aerial photographs and other forms of watershed land use information and human population density

trends can help make extrapolations regarding the system’s response to nutrient loading.

Parameters to Survey

Each of the core variables discussed in Chapter 4 must be included in the survey (e.g., concentrations of

TN, TP [total phosphours], chlorophyll a, and a measure of water clarity, such as Secchi disc,
Nutrient Criteria—Estuarine and Coastal Waters 5-19

submersible PAR meter, or spectral radiometer). It is also appropriate to measure salinity, water

temperature, flow and direction, tide phase, pH, and nutrient load to help better interpret the core

variables. This is a much different problem than usually experienced in rivers and lakes. It may be

desirable in some circumstances to include secondary variables, for example, vertical dissolved oxygen

profiles, distribution and abundance of SAV/seagrasses, distribution of tidal emergent marshes,

distribution and density of benthic filter feeders (e.g., oysters), water color, dissolved organic carbon

(especially if humic-like materials are abundant), and particulate organic carbon. This more complex

array of variables would require a diagnostic justification.

Sampling Frequency

A single grab sample from an estuary or coastal water will be grossly inadequate. Estuaries are near the

bottom of watersheds, which makes them prone to episodic rainfall events. Coastal waters are also

subject to seasonal storms that churn the waters and physically disturb shallow sediments, and these

events may be seasonally highly variable. If information is available to set expectations when possible

seasonal pulses of freshwater occur, then it should be used to help schedule the sampling of wet and dry

periods. In north temperate estuaries, where winter to early spring is the dominant freshet period, this

interval should be included in the sampling scheme. A lag of hours, days, or several weeks to one or

more months is usually required to detect the system’s response to the nutrient load, depending on

magnitude of the freshet relative to volume of the estuary or mixing zones of coastal waters. This also

may capture any spring blooms of diatoms if such occur. A midsummer and early fall survey should give

a first-order picture of the nutrient concentration and response variable pattern suitable for classification. 

In the event of variable summer freshwater flows, then more frequent sampling may be required. 

Because different patterns of rainfall exist around the coasts, regional considerations should weigh

heavily in the design of sampling schedules.

Long-term datasets have well-documented ecological value (Likens 1992, Wolfe et al. 1987, Livingston

2001a); however, all too frequently resources constrain longer term sampling which can average out

short-term variability. Recent data connected to long-term trends provide the strongest case for

classification, reference condition determination, and other criteria development. By measuring the

nutrient load, especially during freshet and low flow periods and concurrently with ambient water quality

and hydrographic sampling, one can get an estimate of the load and salinity/nutrient and response

variable relationships while keeping in mind the precautions noted above. For comparative purposes, it

is important to compare core monitoring variables under similar salinity conditions. 

If tidal elevation is large (e.g., greater than 2.0 m), then this component of estuarine flushing probably

dominates over nontidal gravitational flows (Monbet 1992), and eutrophication symptoms are likely to be

of a small magnitude. In some estuaries (e.g., York River estuary of the Chesapeake Bay), spring tides

may break down density stratification, and the system responds differently to nutrient supplies than

during periods of relatively strong stratification (Hass 1977). Thus, for estuaries with tidal elevations

less than 2.0 m, it is important to note that they are likely to be quite vulnerable to nutrient enrichment. 
5-20 Nutrient Criteria—Estuarine and Coastal Waters

A general rule of thumb regarding freshwater run-off events to estuaries is that a large freshet may

displace the nutrient supply and responses will be detected seaward of the focal area. A modest freshet

may not deliver enough nutrients or physically affect the density stratification to make an estuary

vulnerable to nutrient enrichment. But an intermediate freshet may cause the focal area to receive a

significant nutrient load and establish a strong vertical density gradient so maximum responses will be

detected (e.g., high average chlorophyll a concentrations and minimum Secchi disk readings). This rule

is less easily applied to coastal waters.

Sampling Locations

Sampling locations depend on the size (and especially the length of an estuary), bathymetry, nutrient

source inputs, and hydrography (especially the longitudinal and vertical salinity profiles). In estuaries,

consideration should be given to tidal freshwater, the turbidity maximum (if one is present), mesohaline

and polyhaline regimes, as well as water below zones of density stratification. 

In large tidal freshwater riverine systems, it is important to employ several stations because this portion

of the estuary may “store” a large supply of nutrients that later advect into the saline reach of the estuary

(e.g., the Hudson River system) (Lampman et al. 1999). Enough samples should be taken to detect

nutrient concentration gradients along the salinity gradient from tidal river to the estuary receiving

waters. Typically, this will require from five to seven stations at a minimum. If the estuary is relatively

wide (e.g., lagoonal systems such as Pamlico Sound, NC, or Pensacola Bay, FL) or has large tributary

creeks, then these features may need independent sampling. Where salinity gradients are distinct both

horizontally and vertically, composite sampling may have severe limitations. Depth variability also

should be considered, for example, main channel, shelf samples, and samples in shallow water near

SAV/seagrass meadows or in emergent marsh channels should be included. Emergent marsh creeks

should be sampled in the summer during high and low tides, because high system respiration may cause

hypoxia/anoxia in these tidal creeks that may be largely natural. Where SAV meadows are poorly

developed, resuspension of bottom sediments may be more common and not represented by open channel

samples.

Serious consideration should be given to some replicate sampling within salinity zones to estimate

variability; however, resources may require a broad picture where gradients become equally or more

important than the physical salinity “zones.” In most cases, analytical levels of detection should be a

trivial aspect of data acquisition for reference characterization. This does not free one from application

of good laboratory quality asurance/quality control (QA/QC) practices, which must be maintained with

appropriate blanks, reference samples, and other considerations to standard analytical measurements.

Citizen Monitoring Programs

Citizen monitoring programs have greatly increased, especially since the early 1980's. Where there is

adequate technical oversight either from within the group expertise or from the outside, such monitoring

efforts can play an important role in assessing trends, identifying “hotspots,” and locating likely sources

of nutrients, especially in smaller estuaries where larger research vessels are not required. Many Federal,

State, Tribal, and local agencies assist citizen monitoring efforts, and these agencies contribute to
Nutrient Criteria—Estuarine and Coastal Waters 5-21

training and direction, development, and implementation of QA/QC procedures, act as a data repository;

and perform analyses on environmental samples collected by citizen groups. Citizen monitoring groups

often can provide more frequent observations, such as visiting a gauging station, than can State

personnel. Citizens also can identify those property holders or resource users not following best

management practices or operating within permit limits. See also Volunteer Monitoring Programs,

above.

5.4 QUALITY ASSURANCE/QUALITY CONTROL

The validity and usefulness of data depend on the care with which they were collected, analyzed, and

documented. EPA provides guidance on data QA/QC (U.S. EPA 1998b) to assure the quality of data. 

Factors that should be addressed in a QA/QC plan are briefly described below, but the reader is referred

to published EPA guidance for specifics. The QA/QC plan should state specific goals for each factor and

should describe the methods and protocols used to achieve the goals. The five factors discussed below

are representativeness, completeness, comparability, accuracy, and precision. 

Representativeness

Sampling program design (when, where, and how you sample) should produce samples that are

representative or typical of the environment being described. Sampling designs for developing nutrient

criteria are discussed earlier in this chapter. 

Completeness

Datasets are often incomplete because of spilled samples, faulty equipment, and/or lost field notebooks. 

A QA/QC plan should describe how complete the dataset must be to answer the questions posed (with a

statistical test of given power and confidence) and the precautions being taken to ensure that

completeness. Data collection procedures should document the extent to which these conditions have

been met. Incomplete datasets may not invalidate the collected data, but they may reduce the rigor of

statistical analyses. Therefore, precautions should be taken to ensure data completeness. These

precautions may include collecting extra samples, having backup equipment in the field, installing alarms

on freezers, copying field notebooks after each trip, and/or maintaining duplicate sets of data in two

locations. 

Comparability

To compare data collected under different sampling programs or by different agencies, sampling

protocols and analytical methods must demonstrate comparable data. The most efficient way to produce

comparable data is to use sampling designs and analytical methods that are widely used and accepted

such as Standard Methods for the Examination of Water and Wastewater (APHA, AWWA, WEF, 1998)

and EPA methods manuals. 
5-22 Nutrient Criteria—Estuarine and Coastal Waters

Accuracy

To assess the accuracy of field instruments and analytical equipment, a standard (a sample with a known

value) must be analyzed and the measurement error or bias determined. Internal standards should

periodically be checked with external standards provided by acknowledged sources. At Federal, State,

Tribal, and local government levels, the National Institute of Standards and Technology (NIST) provides

advisory and research services to all agencies by developing, producing, and distributing standard

reference materials. For calibration services and standards see:

http://ts.nist.gov/ts/htdocs/230/233/home/calibration.html. 

Standards and methods of calibration are typically included with CTD sondes, turbidity meters, pH

meters, DO meters, and DO testing kits. USGS, EPA, and some private companies provide reference

standards or QC samples for nutrients. Reference standards for chlorophyll are also available from EPA

and some private companies, although chlorophyll standards are time and temperature sensitive because

they degrade over time.

Variability

Natural variability, rather than imprecision in the method used, is usually the greatest source of error in

the constituent measured. The variability in field measurements and analytical methods should be

demonstrated and documented to identify the source of variability when possible. EPA QA/QC guidance

provides an explanation and protocols for measuring sampling variability (U.S. EPA 1998). Methods for

creating a chlorophyll standard to determine if the spectrophotometer is measuring chlorophyll

consistently from one year to the next or from the beginning to the end of an analytical run are described

in Wetzel and Likens (1991). In addition, replicates for each sample time and site (usually three) must be

collected because the largest source of variation is likely to be natural (i.e., in the samples).

5.5 STATISTICAL ANALYSES

Statistical analyses are used to identify variability in data and to elucidate relationships among sampling

parameters. Several statistical approaches for analyzing data are mentioned here. We advocate simple

descriptive statistics for initial data analyses, that is, calculating the mean, median, mode, ranges, and

standard deviation for each parameter in the system of interest. The National Nutrients Database

discussed above calculates simple descriptive statistics for queried data. Specific recommendations for

setting criteria using frequency distributions are discussed in Chapter 7.

Data Reduction

Data reduction requires a clear idea of the analysis that will be performed and a clear definition of the

sample unit for the analysis. For example, a sample unit might be defined as “an estuary during

July–August.” For each variable measured, a median value then would be estimated for each estuary in

each July–August index period on record. Analyses then are done with the observations (estimated

medians) for each sample unit, not with the raw data. Steps in reducing the data include:
Nutrient Criteria—Estuarine and Coastal Waters 5-23

C Selecting the long-term time period for analysis

C Selecting an index period

C Selecting relevant chemical species

C Identifying the quality of analytical methods

C Identifying the quality of the data recorded

C Estimating values for analysis (mean, median, minimum, maximum) based on the reduction

selected.

Frequency Distributions

Frequency distributions can be used to aid in the setting of criteria. Frequency distributions do not

require prior knowledge of individual waterbody conditions before setting criteria. Criteria are based on

and, in a sense, developed relative to the population of systems in the Region, State, or Tribal

jurisdiction.

Data plotted on a scale of mean nutrient concentration versus frequency of occurrence for a specific

estuary, portion of an estuary, or coastal reach produces a frequency distribution of mean or median

nutrient concentration. Plots of frequency distributions of median TP, median TN, median chlorophyll a,

and Secchi depth for the index period (discussed in Chapter 4) should be examined to determine the

normality of the data in the distribution and to determine the potential for further subdivision of the

waterbody under investigation. Data that are not normally distributed often are transformed into a

distribution more approximating the normal distribution by taking the logarithm of each value. Analysis

of outliers may assist in explaining variability in small datasets; additional analysis can be conducted to

identify the statistical significance of population differences. 

Correlation and Regression Analyses 

The relationship between two variables may be of use in analyzing data for criteria derivation. 

Correlation and regression analyses allow the relationship to be defined in statistical terms. A correlation

coefficient, usually identified as r, can be calculated to quantitatively express the relationship between

two variables. The appropriate correlation coefficient is dependent on the scale of measurement in which

each variable is expressed (whether the distribution of data is continuous or discrete) and whether there is

a linear or nonlinear relationship. Results of correlation analyses may be represented by indicating the

correlation coefficient and represented graphically as a scatter diagram that plots all of the collected data,

not just a measure of central tendency. The statistical significance of a calculated correlation coefficient

can be determined with the t test. The t test is used to determine if there is a true relationship between

two variables. Therefore, the null hypothesis states that there is no correlation between the data variables

measured within the population. A critical " value is chosen as a criterion for determining whether to

reject the null hypothesis. If the null hypothesis is rejected, the alternate hypothesis states that the

correlation at the calculated r value between the two variables is significant. 

Regression analysis provides a means of defining a mathematical relationship between two variables that

permits prediction of one variable if the value of the other variable is known. In contrast to correlation

analysis, there should be a true independent variable (a variable under the control of the experimenter) in
5-24 Nutrient Criteria—Estuarine and Coastal Waters

regression analysis. Regression analysis establishes a relationship between two variables that allows

prediction of the dependent variable (predicted variable) for a given value of an independent variable

(predictor variable). However, scientists (other than statisticians) apply regression analyses to field data

when a relationship is known to exist, even when there is no true independent variable (e.g., cell counts

of algae and chlorophyll concentration; nutrient concentrations and chlorophyll concentration) (Ott 1988,

Campbell 1989, Atlas and Bartha 1993, Ott 1995).

Tests of Significance

Various statistical tests are used to assess the hypotheses being tested. Statistical tests of significance

differ in their applicability to the dataset of interest and the power of the test (the ability of the test to

detect a false null hypothesis). A parametric test of significance assumes a normal distribution of the

population. Nonparametric analyses are valid for any type of distribution (normal, log-normal, etc.) and

can be used if the data distribution is not normal or unknown. A parametric test has more power than a

nonparametric test when its assumptions are satisfied. Two types of errors can be made when testing

hypotheses: Type I—where a correct null hypothesis is mistakenly rejected, and Type II—when there is

a failure to reject a false null hypothesis. The parametric test is less likely than a nonparametric test to

make a Type II error, when the assumptions are met. Therefore, if given a choice, the parametric test

should be used rather than the nonparametric test when the assumptions of the parametric test are

fulfilled. Less powerful, nonparametric tests of significance must be used in cases where the data do not

fit the assumption of a normal distribution (Ott 1988, Campbell 1989, Atlas and Bartha 1993). 

Parametric tests include the student t test, analysis of variance, multivariate analysis of variance, and

multiple range tests. Nonparametric tests include chi square, Mann Whitney U test, and the

Kruskal–Wallis test (Ott 1988; Campbell 1989; Atlas and Bartha 1993). Detailed descriptions of these

and other relevant statistical tests can be found in standard statistical texts.
Nutrient Criteria—Estuarine and Coastal Waters 6-1

Introduction and Definition

Significance of Reference Conditions

Paucity of Similar Estuarine and Coastal Marine Ecosystems

Approaches for Establishing Reference Conditions

CHAPTER 6

Determining the

Reference Condition

6.1 INTRODUCTION AND DEFINITION

A reference condition is the comprehensive representation of data from several similar, minimally

impacted, “natural” sites on a waterbody or from within a similar class of waterbodies, i.e., median

values of TN, TP, chlorophyll a, or Secchi depth. However, in cases where severe degradation has

occurred, surrogate values for reference site data may be required as described in this chapter. There are

two basic approaches for their determination: (1) analysis of in situ estuarine and coastal data, and (2)

analysis of watershed nutrient loading to estuaries and, through advective transport, nutrient loading to

the coastal environment. These approaches reinforce each other, but one may be preferred or even

required depending on comparative or site-specific data. Reference conditions are a primary element of

nutrient criteria development, but should be used in conjunction with the other elements described in

Chapter 1 and Chapter 7. Classification of estuaries and coastal waters should facilitate development of

reference conditions but, as pointed out in Chapter 3, further research is essential to bring classification

of these systems to the level of practical utility comparable to that of most freshwater systems. Models

of estuarine susceptibility to nutrient overenrichment are at an early stage of development, and even less

may be known about coastal ecosystems (Chapter 3).

6.2 SIGNIFICANCE OF REFERENCE CONDITIONS

The reference condition is made explicit through several environmental measures. This manual focuses

on TN and TP as principal causative agents, but their relative roles depend on individual

watershed/estuary and conditions. There are two response variables: chlorophyll a, a measure of algal

biomass; water clarity, linked to algal biomass through chlorophyll a; and often a third, dissolved

oxygen deficiency, particularly in estuaries. These explicit measures are indicators of nutrient

enrichment but are linked conceptually to a continuum of biological resources and recreational

opportunities (Figure 6-1). These linkages show considerable variability and apparent elasticity because

ecological processes are not “hardwired” as are some physically engineered systems (e.g., cogs or pulleys

that drive a machine). States and authorized Tribes are encouraged to employ additional response

variables. 

EPA assumes that ecosystems will support natural assemblages of aquatic life and high-quality

recreational activities if a nutrient supply is achieved and maintained at a level to support the natural

biological system. This can happen, of course, only if other environmental conditions are compatible. 
6-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure 6-1. Environmental quality scale representing reference conditions and potential nutrient criteria relative to

designated uses.

Conditions that support minimal or unimpaired aquatic resources are typically associated with very low

human population densities and limited land use activities in watersheds that would otherwise be a

source of increased human-mediated nutrient supplies to estuaries and coastal waters. This is a plausible

association because once a watershed is moderately to heavily developed, it is practically impossible to

control all nutrient inputs to the aquatic system. State and national preserves may have relatively highquality environmental conditions, but even in these systems atmospheric nitrogen deposition likely has

caused some nutrient enrichment. Ideal conditions may still exist in some estuaries and coastal waters,

and they should be identified. However, where these ideal nutrient conditions no longer exist, especially

in the 50% to 60% of moderately to heavily enriched estuaries (Bricker et al. 1999), the reference

condition should be sought from comparative analyses of similar systems and/or the historical record,

which provides an implied reference condition.

In this manual, a distinction is made for estuaries and coastal waters between “pristine” and minimally

impacted waters. The precolonial period may have had water quality and habitat conditions, including

nutrient loading, that were pristine. This would approximate the ideal of “restoration and protection of

physical chemical and biological integrity” that is the ultimate goal of the Clean Water Act. But this

ideal is largely hypothetical, because methods to estimate it invariably contain a relatively high degree of

uncertainty. However, “fishable and swimmable” conditions are commonly used as an interim goal of

the act and represent a goal of the Nutrient Program, exemplified by reference conditions associated with

minimal human-mediated nutrient enrichment. 
Nutrient Criteria—Estuarine and Coastal Waters 6-3

The term “fishable and swimmable” is not easily quantified in any waters because of their inherent

natural variability. A lake example helps provide a perspective. When a lake shifts from an oligotrophic

to a mesotrophic fishery because of nutrient overenrichment, at some point the change becomes

demonstrable. Likewise, in an estuary when nutrient enrichment shifts phytoplankton production and

algal species composition toward microbial dominance and away from oyster production, the change

becomes demonstrable. What is needed are early warning indicators of the impending change. This is

the role of nutrient criteria. Although it is clear that additional research is called for, there is sufficient

symptomatic knowledge such as changes of species with enrichment to merit setting such indicator

criteria.

Identifying a reference condition in degraded waters should start with an analysis of the best existing

estuarine or marine waters within a watershed or coastal area, or as commonly stated, “the best of what’s

left.” Because of the difficulties in identifying reference sites in some overenriched estuaries and

nearshore coastal waters, it may be necessary to derive an “implied” reference condition by comparing

the “best of what’s left” with “what used to be” as established by the historical record. In any case, it is

still important to identify the best remaining sites in the waterbody of concern. Where the the “best”

sites are known to be severely degraded, more emphasis must be placed on the historical record, but some

knowledge of the continuum from past to present is necessary to establish protective criteria.

Nutrient enrichment-based impacts are a function of the concentrations and supply of nutrients as well as

the ecological conditions and nutrient processes characteristic of the system. Nutrient enrichment effects

may be exacerbated in estuaries where the dominant grazing populations (e.g., menhaden and oysters)

have been lost through human causes, natural causes, or a combination of the two over the past century or

the past several decades. For this reason, it is important to assess the factors that may have modified the

“assimilative capacity” of coastal waterbodies. Reference conditions are not threshold values (a

concentration less than some specified value). For example, in the lower Potomac River estuary,

Chesapeake Bay, the average N load caused a larger phytoplankton bloom than either drought or flood

conditions (Chapter 2). This response involved system hydrodynamics. The point is that reference

conditions should be interpreted in an ecosystem context, especially when the system has experienced

significant nutrient overenrichment and/or is subject to periodic natural disruption such as hurricanes,

winter storms, or droughts.

Some argue that one of the earliest symptoms of impairment involves a nutrient stimulation of harmful

algae relative to beneficial algae. Many types of algal blooms that become a nuisance or harmful clearly

are a form of pollution. There is mounting evidence that many harmful algal blooms are associated with

nutrient enrichment (NRC 2000). For example, substantial loss of seagrass habitat due to algal shading

or bottom habitat due to hypoxia are associated with negative effects on living resources. Seagrasses also

stabilize shorelines and provide cover from predators. 

This manual emphasizes the importance of reference conditions to address nutrient problems in a timely

manner. They serve as the best initial measure for identifying nutrient loads that could cause use

impairments. Statistical and computer-based modeling can improve site-specific estimates of the load
6-4 Nutrient Criteria—Estuarine and Coastal Waters

and response relationships. Classification may assist in extrapolating nutrient effect relationships

between systems in the same class, although classification of coastal waters probably has less

predictability and utility than it does for lakes and streams. As pointed out in Chapter 2, for coastal

waters, the relationships between nutrient loading (e.g., TN and TP) and the response indicators of

chlorophyll a and water clarity normalized to chlorophyll a can be less than straightforward. However,

an understanding of the reference condition will help prevent resource managers from being blindsided

by complications associated with cause-effect relationships.

6.3 PAUCITY OF SIMILAR ESTUARINE AND COASTAL MARINE ECOSYSTEMS

Estuaries and coastal marine ecosystems tend to be relatively individualistic in their sensitivity and

response to nutrient overenrichment. Susceptibility to nutrient enrichment ranks as a premier research

need (Hobbie 2000). The lack of physically similar waterbodies may severely limit grouping

(classifying) waterbodies as recommended for lakes, reservoirs, rivers, and streams where frequency

distributions are used to derive reference conditions (e.g., upper 75th percentile of a priori nutrientunimpaired waterbodies or lower 25th percentile of all waterbodies; EPA 2000a,b). As mentioned in

Chapter 1, an exception may be coastal embayments that form behind barrier islands. If several

relatively similar embayments can be identified within a given geographic area, then one or more may

serve as benchmarks against which the others may be compared.

6.4 APPROACHES FOR ESTABLISHING REFERENCE CONDITIONS

Three primary estuarine approaches may provide considerable flexibility to meet the diversity of

conditions encountered by resource managers. A fourth approach focuses on nutrient loading from the

watershed. A fifth approach is described for coastal marine waters. There are situations where light and

estuarine flushing limit the expression of nutrient enrichment effects. In such cases, downstream effects

may nonetheless be a problem requiring attention. Consequently, we have proposed a variety of criteria

development approaches. Reference conditions in any case should be defined with due consideration of

salinity gradients and seasonal and interannual variability. Table 6-1 summarizes the five approaches to

establishing reference conditions in estuaries and coastal waters.

The alternatives for reference condition determination presented in the following text and table represent

two general approaches to developing baseline nutrient quality measurements. The reference conditions

approach per se acknowledges that the system response to overenrichment can be extremely variable, as

described earlier in this manual. For that reason, measuring the nutrient characteristics of minimally

impacted sites provides a reliable nutrient goal regardless of how those nutrients may or may not be

assimilated. This is the approach upon which the National Nutrient Criteria Program is predicated, and

reference sites should always be sought when designing nutrient criteria protocols.
Nutrient Criteria—Estuarine and Coastal Waters 6-5

Table 6-1. Summary of estuarine and coastal nutrient reference condition determinations

Degree of apparent estuarine degradation Method recommended Criterion measure

A. In Situ Observations as the Basis for

Estuarine Reference Condition

1. Recognized unique excellent condition

2. Some degradation, but reference sites exist

3. Significantly degraded, including all potential

reference sites

Median ambient concentration. Fig. 6-2.

Upper quartile. Fig. 6-2.

Intercept value on a regression or

distribution curve as illustrated in Fig. 6-

3 and 6-4 or by use of a comparable

comparative regression model.

Concentration of TP, TN,

chlorophyll a, Secchi depth

(m)

Same as above

Same as above

B. Watershed-Based Approaches for

Estuarine Reference Condition

4. Same as approach 3 above, but insufficient

historical data

Ref. sites along each trib. and calculate

delivery. Summation is reference

condition. Fig 6-5. Model required to

back-calculate load where all tribs. are

degraded

Load of TP and TN; model

is required to convert load to

estuarine concentration

C. Coastal Reference Condition

5. Applicable to all coastal reaches

 - Estuarine plumes

 - Coastal areas

Index site approach; models may help

distinguish anthropogenic contribution

See also Appendix H.

Concentrations

However, many estuaries are so degraded and/or exhibit such short retention times that investigators

cannot determine reference sites with any degree of confidence. In this instance, dose-response curves

and similar approaches as described below may be more appropriately applied in determining historical

reference conditions, i.e., conditions before degradation was first exhibited. Although these approaches

may not provide the real-time affirmation of existing reference sites, their strength is the documentation

of system decline coincident with the overenrichment. The investigator is expected to assess the

characteristics of the given estuary and select the most responsive option for reference condition

determination from those presented.

In all cases, historical information is important either as an alternative reference condition for heavily

enriched systems or as one of the five elements of criteria development essential to providing a status and

trend perspective important to data evaluation before criteria are established.

In Situ Observations as the Basis for Estuarine Reference Condition

These approaches require nutrient and relevant hydrographic data within an estuary.
6-6 Nutrient Criteria—Estuarine and Coastal Waters

Recognized Unique Excellent Condition

Typically, this condition is based on an extensive spatial and temporal scientific database. If an existing

excellent condition is agreed to by the RTAG and stakeholders, then the State, authorized Tribe, or

appropriate government agency may establish reference conditions based on data that document the

condition and address the remaining elements of criteria development. With limited data, in a very small

number of cases it may be possible to document that the watershed is unimpacted (e.g., has very little

human development, is distant from the influence of local population centers, adjacent land uses are

relatively undisturbed, and is outside of major atmospheric deposition of nitrogen). It is necessary to

document, or augment with additional data, to ensure the original hypothesis is confirmed. Segmenting

the estuary by salinity zones, typically based on estuarine circulation, may be required to reflect nutrient

conditions associated with salinity gradients. However, the geographic scale of the nutrient

overenrichment problem suggests that this a priori approach likely will be limited to a relatively small

number of estuaries and coastal waters (Chapters 1 and 2). The data can be summarized as either

medians of the indicator endpoints or frequency distributions (Figure 6-2). Areas that meet or come very

close to minimally impaired conditions include Plum Island Sound and Blue Hill Bay, ME, and lower

Narragansett Bay, RI. 

Some Degradation Exists But Reference Sites Can Be Identified

Two situations may be applicable for characterizing reference conditions where some degradation occurs.

A. Some Minimally Impaired Sites Are Available

Reference sites should be representative of the system (e.g., a branched estuarine system where one

branch is unimpaired by nutrients and otherwise relatively similar ecological conditions prevail). 

Comparisons should be made among similar salinity zones. This example may apply only rarely because

in atmospheric nitrogen deposition and land use practices likely will have so altered the landscape that

truly undisturbed conditions are unavailable. In those cases where minimal biological resource uses are

impaired by nutrient overenrichment, then reference conditions for nutrients should be deemed to occur. 

Clearly, point and nonpoint source discharges must be at a minimum. Land cover in the watershed

should be very close to natural for the ecoregion or, if modified in the past, then recovery must be well

along (e.g., forests should be near the anticipated climax condition for the region). The reader is referred

to Chapter 4 of “Estuarine and Coastal Marine Waters: Bioassessment and Biocriteria Technical

Guidance,” December 2000 (EPA-822-B-00-024), for additional information (U.S. EPA 2000b). 

Although the upper estuary likely does not qualify, reference-quality sites probably occur in lower

Delaware Bay, especially under sustained periods of low Delaware River input. Under these conditions,

because the measurement values likely will not exhibit a trend or show high variability, the data can be

summarized in terms of median values or a frequency distribution.
Nutrient Criteria—Estuarine and Coastal Waters 6-7

Figure 6-2. Hypothetical frequency distribution of nutrient-related variables showing quantities

for reference or high-quality data and mixed data (all data incuded).

B. Reference Condition Derived From A Priori Selection of a Subset of Reference Estuaries or Coastal

Waters Within a Class of Systems

This approach may apply to a series of coastal embayments located within a physically similar reach of

coastline (e.g., some estuaries along the Maine coast and possibly New England salt ponds). Where

freshwater inputs contribute to a strong salinity gradient, salinity zones may be compared. The goal is to

establish frequency distributions for similar systems. If a population of 15 or more estuaries or

embayments occurs within the class, then a frequency distribution may be applicable (see Chapter 6,

Nutrient Criteria Technical Guidance Manual–Lakes and Reservoirs, April 2000, EPA-822-B00-001). A

sample size of fewer than 15 systems will not likely have enough predictive power to justify application 

of this approach. However, if only a small number exists, then one or two embayments may serve

qualitatively or possibly semiquantitatively as references for the remaining systems. Seasonality and

similar freshwater inputs are important to ensure that physical processes are not masking potential

nutrient effects. To date, it is unknown whether there are enough reference embayments to make a

frequency distribution approach feasible. Other examples may be difficult to identify but tributaries to

North Inlet, SC, may qualify.
1

It is recommended that future algal pigment analyses consider some high-performance liquid chromotography (HPLC)

measurements that can also quantify algal groups.

6-8 Nutrient Criteria—Estuarine and Coastal Waters

Significant Degradation Exists—Reference Sites Cannot Be Identified From Current 

Monitoring Data

Historical Analysis of Estuarine Data

Because of nutrient overenrichment, it is highly likely that no minimally impacted estuarine waters are

available for many coastal areas. Under these circumstances, the “best” of the existing waters are clearly

degraded. An alternative is to establish reference conditions from the historical record. Three

approaches are recommended: (1) analysis of historical ambient nutrient and hydrographic data; (2)

analysis of sediment cores to reveal the historical record, including the paleoecological record; and (3)

model hind-casting. The analysis of ambient data is likely available only for a small number of estuaries

(e.g., Chesapeake Bay and some tributaries, San Francisco Bay, Narragansett Bay, Tampa Bay), because

many estuaries were impaired before nutrient data were initially collected. 

Empirical In Situ Data Analysis

The first approach depends on the availability of an adequate database. Such a database will need

considerable scientific judgment even for systems with a relatively rich database. For example, the

ambient nutrient-based historical record for the entire main-stem Chesapeake Bay becomes much less

abundant during the 1970s and earlier (Harding and Perry 1997). In addition, the spatial coverage was

much less widespread in earlier years. In many systems, early studies focused primarily on the deep

channels because of the interest in hypoxia. Analysis of ambient trends in chlorophyll a may be

complicated by the interaction with freshwater flow, as reported by Harding and Perry (1997) and Hagy

et al. (2000). This co-linear effect may affect assessment of trends in water clarity and cause a nonlinear

relationship between a limiting nutrient and hypoxia (Chapter 2). Hypoxia is the result of an extended

temporal effect of nutrient loading (lag effect), so published empirical relationships usually are based on

seasonal or annual nutrient loads (Chapter 2), not on short-term concentration data. 

An important aspect of the historical approach is the selection of a period when nutrient loading caused

minimal use impairments. Often the nutrient monitoring data are more continuous for past decades than

is documentation of use impairments (e.g., visual pollution and reduction in fish productivity) due to

nutrient overenrichment. The economics and technology that affect fishery statistics in estuaries and

coastal waters are not easily translated into quantitative fish population abundance data. In some systems 

(e.g., Tampa Bay, Janicki and Wade 1996, Greening et al. 1997), seagrasses were monitored to

demonstrate recovery but not necessarily prior to nutrient-based losses. The effects of increased nutrient

loads may be confounded by increased suspended sediments. Sediments can cause light limitation and

impair benthic habitats, with potential negative effects on living resources. 

Historical ambient nutrient-related data will encompass seasonal and interannual variability. It is

important that the key variables (TN, TP, chlorophyll a,

1

 and water clarity—usually Secchi depth) be

distributed in a representative spatial and temporal manner. Although the Secchi disc is inexpensive and

has been widely used, it is recommended that future measurements of water clarity employ submersible
Nutrient Criteria—Estuarine and Coastal Waters 6-9

Figure 6-3. Hypothetical example of load/concentration response of estuarine biota to increased enrichment. 

Dashed line represents the selected reference condition level.

PAR meters as a minimum and that submersible spectral radiometers will be used more frequently. 

Some precautions are relevant. For example, comparing nutrient-related variables in the turbidity

maximum, if light limitation dominates algal net production, would seem to provide little insight into the

nutrient problem. Another consideration is to ascertain where on the long-term hydrograph the reference

period falls (e.g., a wet or dry decade). These past measurements can then be used as a basis for

comparison to present condition (Figure 6-3). If a spring bloom is evident in the data, a judgment is

required as to whether remineralization of the spring bloom helps fuel a summer algal bloom, and

whether the spring bloom contributed directly to summer hypoxia or other use impairments. Therefore, it

is recommended that two averaging periods, a spring and summer period, be used. 

The magnitude and duration of historical algal blooms are expected to be much lower than those of

current blooms in nutrient overenriched waterbodies under similar physical conditions. Some evidence

indicates that the magnitude of algal blooms (e.g., coefficient of variation) may increase as systems

become more enriched. At some point light limitation may limit variability. The historical data should

be aggregated within a physical classification, such as salinity zones, similar to present data analysis. 

Several options exist for data summation. A point-in-time estimate may not capture the anticipated

seasonal and interannual variability in the data. In some cases, data for one or more key variables may

exist before an inflection, indicating worsening conditions. Such variables may include increase in

concentrations or loads of TN or TP, increase in chlorophyll a, decrease in Secchi disk values, and

increase in hypoxic volumes. If this more ideal case prevails, then seasonal medians (medians are less

sensitive to outliers or extreme values in a distribution than are means) or medians of seasonal index

periods (month of highest chlorophyll a) should be calculated. If available, the median of seasonal

medians for one or more years is desirable and may be essential to capture interannual variability.
6-10 Nutrient Criteria—Estuarine and Coastal Waters

Figure 6-4. An illustration of the comparison of past and present nutrient data to establish a reference condition for

intensively degraded estuaries. The option of selecting the distributions from both time periods is compared to an

expected frequency distribution if the observations were available.

Years may need to be excluded from the reference period if the reference period includes increasing

problem-related values (e.g., chlorophyll a and extinction coefficients). Nutrient and chlorophyll a

concentrations and Secchi disk values often will show a gradient within a given salinity zone. In datarich cases within similar salinity zones, one may express the values as a frequency distribution (e.g.,

Figure 6-4; see EPA 2000a, Chapter 6, EPA-822-B00-001, for more details). However, attention needs

to be paid to potential confounding effects (co-linear) of freshwater input, even when comparing similar

salinity zones, because vertical density stratification may be a dominant controlling influence and be

dissimilar over time. When the above precautions are addressed and confounding factors are poorly

understood, it may be appropriate to set the reference condition at the median between the historical

median and the median for present data (Figure 6-4). This simple procedure reflects the magnitude of the

departure from minimally impacted waters and is, in part, a function of the length of the historical

database, addresses inherent variability, and is a realistic approximation of a reference condition over the

time span.

Sediment Core-Based Reconstruction

Sediment core analysis is becoming widely used in assessment of historic biogeochemical and

climatological conditions in lake and marine environments (Brush 1984, 1986; Cooper 1995). The

approach is applicable in sediment depositional areas where bioturbation and other forms of sediment

disturbance are minimal. Improved sediment dating techniques (e.g., lead-210, cesium-154, carbon-14)

have contributed to the understanding of historical conditions when sediments were deposited. Certain 
Nutrient Criteria—Estuarine and Coastal Waters 6-11

other chemicals (e.g., plasticizers) began to leave signatures when developed in the 1930s. Sediment

cores may provide information about metals and organic chemicals with long half-lives. Many

environmental indicators of past sediment conditions now are widely reported in the literature. These

include pyrite formation related to anoxia, history of submerged aquatic vegetation, diatom composition

related to nutrient enrichment, presence and composition of certain dinoflagellate species, distribution

and abundance of indicator species of Foraminifera, hard body parts of molluscs and crustaceans, land

use based on oak and ragweed pollen, total nitrogen and phosphorus profiles, and stable isotope analysis

of carbon, nitrogen, and sulfur in organic materials. Hind-casting with sediment cores provides a means

to infer reference conditions at a time when nutrient concentrations were much lower than present. In

shallow estuaries where depositional areas are not present or questionable, sediment cores have limited

applicability.

Model Hind-Casting

Hind-casting is a controversial approach because it is difficult to calibrate and verify a model when the

data difficult to quantify, as with past nutrient and hydrographic conditions. Running a computer-based

environmental process model backwards involves many uncertainties; however, where ambient data are

inadequate, sediment cores are not applicable, and a model is available and in the hands of experienced

scientists, then with expert guidance it seems reasonable to estimate reference conditions on a first-order

basis. Use of geographic-based land use models coupled to estuarine hydrodynamic-algal growth models

is one approach to hind-casting. Chapter 9 describes several models that may be useful.

Watershed-Based Approaches for Estuarine Reference Condition

An alternative to using in situ coastal marine data to establish reference conditions is to base estimates on

watershed nutrient loading characteristics. One can, in some situations, use watershed loading estimates

to define conditions in the watershed where nutrient loads would represent minimally impaired waters

downstream in estuaries. This assumes that in an undisturbed estuary and its watershed, the nutrient

load would historically represent the most natural condition. In some cases, the watershed above the

confluence of major freshwater streams may contain a relatively low level of human development, but

the nearshore estuarine area may contain considerable development (e.g., Elevenmile Creek, Perdido

Bay, AL/FL, and development near lower Perdido Bay distributary systems; see Livingston 2001a). 

Some estuarine watersheds may still contain a tributary, or segment thereof, whose nutrient load

represents a minimally impaired stream system because human development is minimal. The minimally

impaired stream nutrient loading to head of tide may be used to estimate the nutrient load for the entire

watershed if certain assumptions are met. Application of a watershed stream segment to estimate

nutrient loading is based on seven assumptions, as described in Table 6-2, including atmospheric

deposition of nitrogen as a constant.

Areal Load Approach to Identification of Reference Condition

The nutrient load is measured for the minimally disturbed subtributary or segment. State and national

preserves, when available, typically offer appropriate conditions (Clark et al. 2000, Fulmer and Cooke

1990). A decision is required as to whether the geology is approximately homogeneous across the

watershed to extrapolate from the reference tributary to the entire suite of tributaries. If not, then a 
6-12 Nutrient Criteria—Estuarine and Coastal Waters

 Table 6-2. Requisite assumptions for establishing watershed-based reference conditions

Assumptions Description

Estuarine systems are usually unique rather

than being readily divided into similar classes of

estuaries.

There are some instances of coastal bays that can be classified together 

so that one or more of them may be designated as reference sites that

collectively may comprise a reference condition against which other

similar estuaries in a given area can be compared. However, most

often each estuary must be addressed individually and the reference

condition must be derived from data within that system.

The tidal factor and large volume shifts make

regional subdivision of each overenriched

estuary difficult.

In the instance of biological criteria development, it was possible to

subdivide the estuary by salinity habitat regimes and thus reduce the

portions of the estuary and its water column dynamics to a manageable

level, especially for assessments of nonmobile benthic invertebrates. 

This is not viable with nutrient-related planktonic organisms and

dissolved or suspended water column materials.

We can estimate “pristine” or natural loads

from estimations of concentration-flow 

relationships, and therefore loading estimations

in unaltered subestuarine watersheds. That

loading can be extended to an entire estuary for

reference condition development. 

Loading estimations are an established practice in water resource

management. The universal soil loss equation (USLE) is the best

known example, and it has been extended to other unit loading

estimations. This robust and rather straightforward concept estimates

most of the nutrients that enter an estuary regardless of how effectively

the system processes or assimilates nutrients. 

Most of the measurable load to an estuary is

tributary based, and atmospheric deposition is a

constant over the system.

Admittedly, there is an atmospheric source of nitrogen to and from

estuarine waters, but this is variable and, generally, is much less than 

runoff from streams and rivers. The effect is somewhat mitigated

because the atmospheric deposition to the stream surfaces is

incorporated in the loading estimation. Phosphorus and suspended

material as well as algal responses to nutrients are most certainly more

tributary related than atmospheric. Shoreline sheet runoff can be

incorporated into the loading approach. 

 

Loading from coastal marine waters is usually

negligible compared with anthropogenic

watershed loads to the estuary

While many estuarine ecologists are properly concerned with this

aspect of estuarine dynamics, from the practical criteria development

approach, a large portion of the marine load may be presumed to have

originated in the previous outgoing estuarine tidal water. The

remainder is to some extent a part of a natural process inherent to

estuarine systems. This assumption may not prevail when estuaries

enter deep, upwelling, oceanic waters. 

The predevelopment nutrient loading rates

expressed as yield per watershed land area are

similar within a single geographic region (e.g.,

province, ecoregion, or subecoregion). Local

regional uniformity of geography is assumed.

This is a reasonable expectation in the absence of extensive land

development with attendant anthropogenic discharges and runoff. The

geographic subdivisions of a natural landscape can be expected to be

homogeneous by definition (i.e., similar soil type, topography, and

vegetative cover).

Because the National Nutrient Criteria Program 

assumes that groundwater influence is a

separate loading factor to surface water

eutrophication, groundwater-dominated

estuaries should be treated separately for the

development of nutrient criteria.

The groundwater factor can be highly significant in some localities

such as coastal Florida. This generalized approach does not address

that factor, but Regional Technical Assistance Groups should be aware

of the groundwater contribution and account for it in their estimations. 
Nutrient Criteria—Estuarine and Coastal Waters 6-13

second minimally disturbed subtributary should be sought to represent a different geology; this logic

should be applied if additional subwatersheds differ substantially. Nutrient loads are then extrapolated in

a simple proportional manner from the reference tributaries to the entire tributary system within the

watershed or subunit and the aggregate load is calculated (Figure 6-5). This load represents the

“reference load” reflecting in situ reference conditions for the four primary indicators (TN, TP,

chlorophyll a, and water clarity).

Extrapolation from a reference tributary can be augmented by application of geographic-based nutrient

erosion and transport models. This nutrient load would become the target load for the downstream

estuary or coastal waterbody. Consideration should be given to the representative nature of the

freshwater flow over the average hydrograph. It is desirable and may be necessary to obtain a measure of

the average multidecadal freshwater flows at the head of tide. Nutrient loads based on a drought period

would not accurately represent conditions in terms of nutrient-based ecological impairments. Extremely

high flows are important, but they are likely to fall outside of resource managers’ capabilities to solve a

nutrient problem. Therefore, this approach has its limits. Large fluvial streams do not necessarily

transport the most upstream load to the lowest fluvial portion of the stream tributary; in-stream

ecosystem processes modulate the load. In summary, a sequence of steps is outlined in the following box

to complete the areal load watershed approach for reference condition determination.

Coastal Reference Conditions

Following is an example of a nutrient loading assessment from a very large watershed, the Mississippi

River system. In 1997, the EPA Gulf of Mexico Program, through a Mississippi River/Gulf of Mexico

Watershed Nutrient Task Force, asked the White House Office of Science and Technology Policy to

conduct a scientific assessment of the causes and consequences of Gulf hypoxia through its Committee

on Environment and Natural Resources (CENR). The National Oceanic and Atmospheric Administration

(NOAA) was asked to lead the assessment. The assessment included various computer-based modeling

approaches to characterize the nutrient delivery to coastal Louisiana and Texas.

Nutrient load reduction within the various watersheds to meet resource management objectives is an

analogue of the reference condition approach (see NOAA website for details:

www.nos.noaa.gov/products/pubs_hypox.html). Two conditions are considered: coastal estuarine

plumes and shelf waters.

Coastal Estuarine Plumes

The foregoing example for the Mississippi River watershed is a large-scale effort to assess nutrient

conditions associated with a large coastal nutrient plume. For large estuarine watersheds with the

potential to cause nutrient overenrichment on the continental shelf, it is important to extend the reference

concept beyond the local area. In many cases, the concerns will be large enough to warrant development

of a hydrodynamic model with coupled nutrient-phytoplankton growth kinetics (Chapter 9). Smaller

estuarine plumes along the coast may be addressed through a well-designed research and monitoring plan

with expert input on design features. An initial EPA-sponsored sampling effort in the Mid-Atlantic is

currently underway to provide range-finding data to assist in development of a more comprehensive
6-14 Nutrient Criteria—Estuarine and Coastal Waters

Figure 6-5. Areal load estimate approach to nutrient reference

condition determination. The watershed is estimated to be

approximately 368 square miles; the reference tributary streams

representative of above head of tide systems in the watershed are

approximately 20 square miles combined. The median load estimate at

the mouths of the tributaries could therefore be multiplied by 18.4 to

approximate a reference condition load for the river.

monitoring plan. The first step is to assess the shelf’s mixing and dispersive capabilities to attenuate the

negative effects of nutrient enrichment. Bloom development along physical discontinuities should be

assessed. If dispersive mechanisms are large enough to attenuate phytoplankton blooms, then concerns

are given a lower priority. In coastal estuarine plumes where the physical processes may not attenuate

the nutrient enrichment effects to acceptable levels, then an appropriate level of research, monitoring,

and modeling may be required to assess nutrient reduction from upstream nutrient sources as well as

from seaward upwelling of nutrients and atmospheric deposition (Paerl and Whitall 1999, Vitousek et al.

1997, Howarth et al. 1996).

Open Coastal Water

The NRC (2000) publication (Chapter 6) recommends an index site approach for estuaries and coastal

waters. The index site approach in particular has merit because the continental shelf is very extensive and

too large for the Nation to conduct comprehensive studies of all sites potentially affected by nutrient

enrichment. A priority-setting rationale should be based on a physical classification system that arrays
Nutrient Criteria—Estuarine and Coastal Waters 6-15

coastal water masses relative to their susceptibility to nutrient enrichment, especially waters that are

likely to respond similarly to nutrient enrichment. This would allow resource managers an opportunity to

apply unique “dose-response” curves to a particular coastal water class. Because such a system likely is

relatively crude, the best professional judgment of experts should augment physical classification

systems. Because this approach has a significant research element, it would be appropriate to begin it in

the competitive research forum with oversight provided by some mix of Federal, State, and university

participation. Chapter 7 also discusses the roles of monitoring and modeling, which offer useful insights

applicable to the coastal ocean. This manual recommends that the index site approach be given serious

consideration, especially for coastal waters including estuarine plumes. The coastal ocean is large and

oceanographically dynamic and complex. Thus, assessment of individual States’ contribution to nutrient

overenrichment will in most cases require a Federal and States’ partnership. Even in large estuaries

whose watersheds or tidal waters are shared by more than one State, a multi-State agreement is probably

required (e.g., Chesapeake Bay Program and the New York and Connecticut TMDL Agreement).

The Coastal Research and Monitoring Strategy, an element of the Clean Water Action Plan, contains

approaches that can help determine reference conditions. The strategy provides for a coordinated effort

among Federal, State, and private agencies. Clearly, an approach that coordinates use of aerial

surveillance tools (e.g., satellite-based water quality sensors), data buoys, and ship-based measurements

(especially ships of opportunity such as the North Carolina Albemarle Sound ferryboat monitoring

program; H. Paerl, personal communication) within an index site to underpin a cause-and-effect

framework is highly desirable. 

On a provisional basis, any additional monitoring might include a stratified random approach (e.g.,

EMAP), because this provides an opportunity to address known ecological structure and functional

processes and unbiased trend monitoring. It is important to continue monitoring that involves identified

relationships. A challenge will be to design a program that can distinguish the effects of natural coastal

processes (e.g., nutrient upwelling) from anthropogenic influences (atmospheric nitrogen deposition,

fluxes of nutrient from estuaries, and possible expansion of mariculture activities).

One such investigation is the design presently being tested by the National Nutrient Criteria Program in

the Mid-Atlantic Bight. This near-coastal marine nutrient sampling protocol is intended to identify

inshore (within the 3-mile limit) reference sites based on land use and physical coastal characteristics

together with comparisons to offshore nutrient water quality. A stratified-random approach is used, and

the compiled data from reference sites establish the reference condition for that portion of the coastal

marine waters. Riverine and estuarine plumes or other discharges can then be evaluated relative to this

minimally impacted condition. Sampling is recommended for spring and summer conditions with

multidepth collections. The technique is being tested in a variety of State waters. A description of the

design and preliminary results is presented in Appendix H.
6-16 Nutrient Criteria—Estuarine and Coastal Waters

1. Identify major tributaries to the estuary. Classify similar tributaries by physical size and freshwater delivery and

compare similar geological features of the subwatersheds for classification.

Three nutrient loadings can be estimated:

A. Existing (actual current load). Existing load is estimated from the known tributary loads as measured at the mouth

above head of tide and extrapolated to similar land units in the watershed, plus any shoreside runoff and discharges

that directly enter the estuary, plus direct atmospheric deposition. This is the present status of the waterbody. 

B. Best existing load. Find the best existing tributary(ies) or best subtributary(ies) in the region and calculate the

nutrient yield. Extrapolate this value to the rest of the watershed tributaries or subtributaries as appropriate.

C. “Pristine” or unimpacted load. Identify a regional reference streams in an undisturbed watershed having little or

no development, such as State or national preserves, that can be used to estimate areal yield for the region

approximating an entire unimpaired, undeveloped condition as though there were no significant cultural impairments.

2. Find the tributary(ies) with the least impaired status and minimal disturbed lands contributing to nutrient

loads. 

These are B and C-type conditions from above. Each of these tributaries is monitored enough to establish a nutrient

load. Seasonal and interannual variability should be assessed. Cross-sectional sampling at the head of tide is required

because of nutrient flux variability. The USGS has established protocols for this type of stream monitoring. Where

available, dams on rivers, if located near the head of tide (e.g., Conwingo Dam on the lower Susquehanna River, MD),

make very desirable locations to measure nutrient loads.

3. Estimate the annual areal nutrient yield for TN and TP.

Extrapolate from the unimpaired or minimally impaired watersheds above head of tide to all other similar tributaries in

the watershed and apply the load estimate to direct runoff portions of the tributary. Do this for each monthly or

seasonal increment throughout the year. Do the same for each major tributary.

4. Extrapolate the nutrient yield to the entire watershed land area within the region.

For example, generate total best existing nutrient loading. This will, in many cases, represent the best attainable loads.

5. Repeat for other regions if the estuary watershed covers more than one major geological landform.

This is necessary to comply with the assumption that regional homogeneity within the watershed covers only part of

the entire estuarine watershed.

6. Sum the nutrient yield for all tributaries within the estuarine watershed.

Do not factor in atmospheric nitrogen loads, as they were incorporated into the tributary loads. The atmospheric

nitrogen loads may need to be reduced to achieve an acceptable nutrient condition in the estuary. The summation of

tributary loads becomes the estuarine reference condition. Computer modeling may be required, especially in larger

watersheds, because in very long tributaries some nutrients, especially phosphorus, may become embedded in the

stream bottom, and some nitrogen and phosphorus is potentially lost on the scale of years to decades in the floodplain. 

The chlorophyll a concentrations in the estuary will need to be modeled from the reference nutrient loads and some

measure of central tendency of freshwater inflows to the estuary. When a larger estuarine system may dominate the

lower tributary estuarine hydrodynamics (e.g., the mainstem of the Chesapeake Bay dominates the hydrodynamics in

the lower Patuxent River estuary), then as a minimum, a box modeling approach may be required to account for the

two-estuarine interactions of the freshwater inflow and lower estuarine interactions (Hagy et al. 2000).
Nutrient Criteria—Estuarine and Coastal Waters 7-1

Role of RTAGs

Classification

Descriptive Background Information

Elements of Nutrient Criteria

Hypothetical Examples of Nutrient Criteria Development

Determinations

Evaluation of Proposed Criteria

Nutrient Criteria Interpretation Procedures

Criteria Modifications

EPA, State, or Tribe Responsibility Under the CWA

Implementation of Nutrient Criteria

CHAPTER 7

Nutrient and Algal 

Criteria Development

7.1 INTRODUCTION

This chapter addresses the details of developing scientifically defensible quantitative criteria for

nutrients, algae, and measures of water clarity (referred to hereafter as nutrient criteria). The chapter is

divided into eight sections: (1) role of the RTAG; (2) classification; (3) descriptive background

information; (4) approaches to criteria development; (5) evaluation of proposed criteria; (6) criteria

modification; (7) EPA, State, and/or Tribal responsibilities under the Clean Water Act (CWA); and (8)

implementation of nutrient criteria into water quality standards. The five elements of criteria

development described in the Executive Summary and Chapter 1 are integrated herein, and additional

information relevant to criteria development is provided.

As explained in Chapters 2 and 3, estuaries and coastal waters are especially complex ecosystems where

it is often difficult to distinguish the effects of anthropogenic nutrient enrichment from natural variability

primarily because of intervening physical processes and their interaction with biological (e.g., grazing;

Cloern 1982) and chemical (e.g., flocculation and sedimentation) processes (Malone et al. 1996, NRC

2000). The individual nature of many of these ecosystems presents a particular challenge for criteria

development. The ideal goal is to establish nutrient criteria that are protective during periods when

estuaries and coastal ecosystems are most vulnerable to nutrient enrichment and that protect designated

uses. It is important to understand that designated uses may be met but some nutrient-based impairment

may have occurred. In such cases, it is also desirable to have restoration goals in mind whose objective

is to restore the original ecological integrity, at least as represented by the reference conditions and

criteria. This information helps determine if new designated uses are appropriate.

If a shallow estuary is dominated by point sources of nutrients, then low freshwater flow periods might

be times of greatest vulnerability because of limited flushing. For a deep estuary under the same

situation, weak density stratification may set up conditions where the algae spend considerable time

below the euphotic zone and, hence, bloom development is minimized. Now, consider the situation

where these two estuaries are dominated by nonpoint sources of nutrients. The low flow period would

likely contribute lower nutrient loads to both types of estuaries and with weaker density stratification the

algal bloom potential might be substantially lowered (see Chapter 2). It is also extremely difficult to

establish nutrient criteria for episodic events, either hurricanes or drought periods. The reference

condition and consistent reference site records, however, make it possible to adjust the criteria

accordingly to address these intervals.
7-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure 7-1. Generalized progression and relationship of the elements of a nutrient criterion.

In this manual, the elements of criteria development are sequenced and emphasized somewhat differently

from those described in the two published freshwater manuals. Because of the relatively high individual

nature of estuaries and coastal waters, the role of the regional technical assistance groups (RTAGs) is

enlarged and historical data become especially important in development of nutrient criteria. Figure 7-1

provides a visual perspective of the elements that should be integrated to arrive at a criterion.

7.2 ROLE OF REGIONAL TECHNICAL ASSISTANCE GROUPS 

Expert evaluations are important throughout the criteria development process. The role of the RTAG in

criteria development for estuaries and coastal waters has an added dimension over that applicable to

lakes and reservoirs and rivers and streams. In the latter case, most of the data used in development of

criteria recommendations resided in national electronic data sets (e.g., STORET) collected by State and

Tribal agencies. The RTAGs helped review these data for duplication and outliers. They also

encouraged States and authorized Tribes to submit additional data to STORET. Under these

circumstances, EPA developed criteria recommendations based on frequency distributions (U.S. EPA

2000a,b). However, for most estuaries and coastal waters, the majority of relevant data have been

collected by universities and other organizations (e.g., NOAA, Minerals Management Service, USGS);

many of their data are not entered into STORET. It is expected that regional RTAGs will likely be

considerably more knowledgeable about the data veracity and applicability for criteria development in

their region. In addition, it is anticipated that the regional RTAGs will have the knowledge and access to

other local scientific expertise to assist in the sampling design and collection of additional data. These

considerations lead to the expectation that the RTAGs will have a larger role in the development of

protective nutrient criteria for estuarine and coastal waters.
Nutrient Criteria—Estuarine and Coastal Waters 7-3

Because of the tendency of estuaries and coastal waters to exhibit a high degree of individuality relative

to nutrient susceptibility, little predictive success has been demonstrated in using published values of

nutrient concentrations or nutrient impairments. However, experience with an individual estuary or

coastal water system may demonstrate a general range of algal biomass accumulation that leads to

hypoxia at a known level of enrichment. This type of assessment often requires several to many years of

observation and measurement and is an example of incorporating the RTAG experience in the criteria

development process.

EPA expects to continue to lead the effort to identify potential estuarine and coastal waters for the

development of nutrient criteria and fund the overall data collection and analysis. 

7.3 CLASSIFICATION

Classification is a pivotal step in the process of developing criteria. Physical classification of

waterbodies for nutrient criteria development reduces variability in ambient measurements as the

reference conditions and nutrient criteria represent and reflect a relatively similar natural state. 

Classification helps ensure that appropriate comparisons are made among comparable waterbodies so

that the variables measured are influenced as little as possible by dissimilar inherent characteristics. This

will facilitate appropriate application of criteria through their implementation. 

In contrast to rivers and lakes, physical classification of estuarine and coastal waters is scale-sensitive

(Giller et al. 1994) and may not be as predictable of nutrient enrichment effects or be as useful for

generalizations about effects among estuarine systems. Classification nonetheless can provide improved

understanding of the processes that contribute to ecosystem susceptibility and variability in the

expression of nutrient effects. Classification may have valuable applicability at smaller physical scales

within larger estuarine and coastal ecosystems (e.g., embayments; subestuaries and estuaries discharge

plumes). Classification based on salinity gradients, circulation patterns, depth, and flushing within larger

estuarine and coastal systems should also prove useful, especially in correlating the different biological

communities at risk to nutrient overenrichment.

7.4 DESCRIPTIVE BACKGROUND INFORMATION

Estuarine Watershed Characterization

One of the keys to understanding nutrient enrichment problems in waterbodies is an environmental

characterization of the watershed from a historical perspective. Such investigations may provide insights

into the potential for confounding cause and effect relationships (e.g., nutrient, not herbicides as the

primary cause of the SAV decline in Chesapeake Bay.) Historical changes in land use, land cover, and

population demographics may correlate with increased anthropogenic-based nutrient enrichment and

major changes in downstream water quality and biological community structure. 

Farm acreage, crop types, and fertilizer application rates also provide useful information to assess the

potential historical magnitude in nonpoint source nutrient loading to estuarine systems. Atmospheric
7-4 Nutrient Criteria—Estuarine and Coastal Waters

plumes that result in nitrogen deposition should also be assessed as they may help explain an increased

nitrogen load to coastal waters beyond that attributed to local land use activities. This also is a situation

where reference sites, if available, will help the manager distinguish atmospheric from local

anthropogenic causes of overenrichment. Conversely, increases in forest coverage and soil banking of

agricultural lands may help explain potential decreases in nitrogen and phosphorus loading to estuaries

and coastal waters.

Within Estuarine System Characterization

Changes within estuarine systems that influence basic hydrography should not be overlooked. For

example, opening new passes, and later deepening them in several Gulf of Mexico estuaries during the

early part of the 20th century (e.g., Perdido Bay, Alabama/Florida, and Choctahatchie Bay, Florida)

apparently modified estuarine circulation resulting in strengthened density stratification leading to

enhanced potential for hypoxia (Livingston 2001a). Also, major changes in freshwater supplies should

be considered as a potential factor that can modify estuarine susceptibility to nutrient enrichment.

The use of marine sediment cores is another tool to assist in assessment of nutrient enrichment patterns in

coastal waters (Brush 1984). These analyses are relatively expensive to perform but appear more

frequently in the literature because of the numerous important insights they provide. They can provide

estimates of sedimentation rates and initiation of anoxia, changes in algal community structure, initiation

of the loss of SAV and other responses to nutrient enrichment. This temporal picture is important in

setting approximate timelines when nutrient enrichment may have been a major cause of biological

impairment. However, correlation does not necessarily equate to causality (Havens 1999).

It is also important to attempt to collect long-term (e.g., multiple decades) fisheries landings data as many

stakeholders need to be appraised of whether such landings data are associated with nutrient enrichment. 

Such analyses often prove to be difficult because information on catch per unit effort that helps

normalize for variable fishing pressure, is difficult to obtain. A change in fishery yield may be

confounded by overfishing, as well as the role of increased or decreased primary productivity. Increased

bottom water hypoxia related to overenrichment may explain the loss of benthic habitat for bottomdwelling marine life (e.g., flounder and croaker and benthic infauna that serve as fish food). 

Historical decreases in water column visibility from nutrient-driven algal blooms (phytoplankton and

macroalgae) may explain reductions in water-borne recreation (e.g., swimming); however, human

perceptional responses are often subjective and other factors such as user conflicts may also be involved. 

Reduced visibility may also be related more to inorganic suspended sediments then nutrients. So, both

chlorophyll a records and total suspended sediment concentrations, preferably measured from the same

local water mass, may be required to establish nutrient enrichment (Gallegos 1994). Such analyses need

to be assessed with freshwater flow records because of potential co-linear effects that complicate

interpretation of cause and effect relationships. 

The above measurements performed as part of a general characterization may help relate nutrient

enrichment effects and thresholds to existing and designated uses and identify any reference systems that
Nutrient Criteria—Estuarine and Coastal Waters 7-5

are minimally impacted by nutrient pollution. Such information may enter directly into predictive water

quality models or serve as indirect collaborative information contributing to a “weight of evidence”

analysis.

7.5 ELEMENTS OF NUTRIENT CRITERIA

Reference Condition

Chapter 6, beginning with Table 6-1, describes several approaches that can be employed to determine the

ambient minimally impacted nutrient condition of the water resource. The significance of this reference

condition to nutrient criteria development cannot be overstated. It represents the determination of the

existing, presently attainable nutrient water quality of the estuarine or coastal waters of concern.

The selection of the method for reference condition determination in this manual involves more options

than previous guidance manuals because so many estuaries are reported to be unique and/or severely

degraded, thus requiring an array of alternative approaches to approximate reference conditions in the

absence of acceptable reference sites. In selecting from the different approaches, the resource manager

should strive for the most direct measurement of the resource and with the least number of intermediate

interpretative steps to a determination. Of equal importance in this process are in situ reference sites and

supporting data showing the system response to the nutrient increases. The best of both these worlds is a

set of reference sites documenting an optimal nutrient condition as well as response data confirming that

system degradation occurs at levels beyond this measure, which also corresponds to the EPA regional

reference condition for that area and class of waters. Failing this, the manager should seek the greatest

approximation possible and a sufficient understanding of the divergence to be confident of the reference

values determined.

Even though the reference condition is salient to nutrient criteria development, it should not be

interpreted as the only necessary element. It should be interpreted in light of the historical condition of

the resource and projections of its future potential.

Historical Information

Knowledge of antecedent conditions is particularly important in the case of estuarine waters, where

causal relationships are often confounding and existing reference sites compromised. In such cases the

historical data may not only qualify present information, they may, in fact, be the requisite reference

condition demonstrating to the manager and RTAG not only previous nutrient quality, but least impacted

conditions as well.

Models

In addition to computer modeling to help determine reference conditions, models may be used both to

estimate nutrient loads and load reductions to achieve a targeted nutrient regime in the receiving

estuarine and coastal waters systems. Various models are available (Chapter 9) that estimate nutrient

erosion from different land uses, riverine transport of nutrients, and estuarine effects including hypoxia

and chlorophyll a concentrations. These models are typically expensive to calibrate and verify, but in
7-6 Nutrient Criteria—Estuarine and Coastal Waters

large estuaries and coastal ecosystems their application can be cost-effective when costly nutrient control

decisions are involved. A coupled nutrient transport and hydrodynamically coupled algal nutrient uptake

and growth model provides the ability to address “what if scenarios” of nutrient load versus level of

biological impairments (see Chesapeake Bay Case Study and web site). 

Statistical models can be used to help separate effects of nutrient loading from estuarine physical

processes as determinants of increases in system response variables (e.g., chlorophyll a concentrations;

Harding and Perry 1997). Effects of nutrient enrichment are inferred by a process of elimination when

the suspected physical forcing functions are de-trended in the analyses and are inferred not to explain the

variability. Box models, using salinity as a tracer of water masses, are useful in assessing net non-tidal

physical circulation, hydraulic residence times, the effect of river flow on residence time versus seaward

higher salinity processes, and extent that nutrient sources are conserved within an estuarine region or

transported seaward (Hagy et al. 2000). Dettmann (in press) used a regression modeling technique to

compare the degree of nitrogen export of a variety of estuarine systems. Boynton et al. (1996) used a box

model to mass balance nitrogen and phosphorus in Chesapeake Bay and calculated net transport from the

Bay of nitrogen and phosphorus, sedimentation, amount tied up in plants and the amount of nitrogen lost

from the Bay through denitrification. Properly applied box and regression models are relatively

inexpensive to construct and can provide useful information to the scientist and water quality manager.

Antidegradation Policy and Attention to Downstream Effects

A critical requirement for the use of reference conditions associated with nutrient criteria is the EPA

antidegradation policy, which protects against incremental deterioration of waterbodies and reference

conditions. An observed downward trend in the conditions of reference sites cannot be used to justify

relaxing reference expectations, reference conditions, and the associated nutrient criteria. Once

established, nutrient criteria should only be refined in a positive direction in response to improved

conditions. Without antidegradation safeguards, even the establishment of reference conditions and

nutrient criteria could still allow for continual deterioration of water quality.

To combat this, the States should implement an effective antidegradation policy that promotes

continually improving conditions. As an example, Maine has an antidegradation policy that requires that

waterbodies remain stable or improve in trophic state (Courtemanch et al. 1989, NALMS 1992). The

RTAG should assume a comparable sense of antidegradation responsibility.

Estuaries that supply nutrients to relatively static coastal waters may require more stringent nutrient

criteria, not only to protect estuarine designated uses (e.g., “fishable and swimmable” conditions), but the

water quality of coastal shelf waters. At present, there are little data to assess whether U.S. estuaries are

supplying nutrients to coastal shelf waters at levels that are causing widespread harm, except in the case

of world-class rivers (e.g. Mississippi River Plume on LA/TX shelf). Locally, river-dominated estuaries

with open passes to the coastal shelf supply nutrients at levels that may increase secondary productivity

of valued fisheries (Sutcliffe et al. 1977), but the potential threshold for overenrichment effects in such

cases is generally still poorly understood. If the RTAG determines that estuarine nutrient criteria may be

expected to fall between the existing present nutrient concentrations or load and the reference condition
Nutrient Criteria—Estuarine and Coastal Waters 7-7

determined from similar unimpaired systems or from a historical load and response relationship. It is

then up to the States and Tribes to adopt the criteria into 303(c) water quality standards.

The RTAG

Assimilation of all of the above information is the responsibility of the RTAG when developing

ecoregional nutrient criteria and when reviewing State or Tribal nutrient criteria as part of its role to

assist EPA.

The RTAG should work with the States to develop a monitoring program that would evaluate the status

of the reference systems, the possible future negative anthropogenic nutrient effects, and the condition of

the estuary and coastal receiving water. A well-designed monitoring program should provide data to

assess whether there is incremental deterioration of the subject waterbodies and reference conditions.

7.6 HYPOTHETICAL EXAMPLES OF NUTRIENT CRITERIA DEVELOPMENT

DELIBERATIONS

To help illustrate the role and responsibility of the RTAGs, an abbreviated hypothetical 

illustration of nutrient criteria development follows for a river-dominated estuary that has a relatively

deep channel with moderate density stratification, and well-developed seagrass meadows located in the

shallow waters. The estuary is located in the northern Gulf of Mexico. The estuary often borders on

both nitrogen and phosphorus limitation with nitrogen limitation occurring more frequently during the

summer. The nitrogen sources have been and continue to be primarily nonpoint sources in headwater

streams and point sources near the pass to the Gulf. Tidal action is minimal. The focus for this

illustration is on nitrogen criteria.

Scenario

In the subject nitrogen-limited estuary no existing areas qualified as a reference condition and no

meaningful analogs of the estuary were available to apply the spatial-based frequency/percentile

approach to reference conditions applied to lakes (e.g., the 25/75 percentile approach; see Lakes and

Reservoirs Nutrient Guidance Manual, U.S. EPA 2000a). The historical nutrient concentrations were

plotted over time by classifying the summer estuarine salinity zones based on the 30-year average into

tidal fresh and brackish (0-5 psu), mesohaline (5-18 psu) and polyhaline (18-30 psu). The concentrations

were consistent with a calculated nutrient load and estuarine physical hydrodynamic model. Additional

analysis demonstrated that the system was most vulnerable to average summer freshwater inflows. 

Flushing in the estuary was determined to be on the order of one month on average during the summer

under average freshwater flow conditions so the physical potential was high for phytoplankton bloom

buildup. 

A 50 μM TN reference condition was determined by plotting TN concentrations for the mesohaline zone

from 1970 to 2000. Because nutrient data were few for the estuary prior to 1970 and most of the data

were collected for the mesohaline zone, a decision was made to select the 1970 summer average TN

value as the reference condition with application for the mesohaline zone. Only an occasional bottom
7-8 Nutrient Criteria—Estuarine and Coastal Waters

Figure 7-2. Hypotetical illustration of developing a TN criterion in an estuary.

channel hypoxia occurred in the summer of 1970 and seagrass meadows were growing to a depth of 1 to

2 meters. By 1980, the average summer TN concentrations had increased to 60 μM when it became

apparent from field monitoring that major loss of seagrass acreage was documented, hypoxia volumes

had doubled, and hypoxia-laden water had consistently reached the deep-channel/shallow water shelf

break. By 1980, seagrasses grew on average only to 0.5 meter depth. Some sediment core evidence

indicated that seagrasses in 1950 grew at a depth of 2 to 3 meters but the nutrient concentrations and

loading information were much weaker than for the 1970-2000 period. The reference condition of 50

μM TN suggests a gradual rise upward in ambient TN, from an estimated average 20 μM TN 100 years

ago, with a projected further upward trend indicated by use of demographic, land use, and hydrological

models. It was determined that a significant loss of ecological integrity had likely occurred prior to 1950. 

The RTAG, therefore, concluded that setting a criterion any higher than the present reference condition

would eventually lead to an unacceptable trend in water quality degradation due to expected development

increases for this part of the estuary. If the model projections are accurate, any increased load implied by

raising the criteria above reference levels will hasten nutrient overenrichment problems. Some on the

RTAG argued that the criterion should be set at 54 μM TN but the final consensus was to be somewhat

conservative. The RTAG therefore concludes that it will be prudent to set the criterion at 50 μM TN (see

Figure 7-2).

7.7 EVALUATION OF PROPOSED CRITERIA

The RTAG will provide expert assessment of proposed criteria and assure that criteria protect designated

uses. Criteria will need to be verified in many cases by comparing criteria that apply across State and

Tribal borders. In addition, attention will need to be paid to downstream effects and designated uses,
Nutrient Criteria—Estuarine and Coastal Waters 7-9

especially in large estuaries that are shared by two or more States or Tribes. Criteria recommended by

the RTAG can be adopted by the State or Tribe and approved by EPA if evidence is presented that

assures no adverse effects will result downstream (e.g., criteria developed for tidal freshwaters may not

be stringent enough to protect uses located at higher salinities.) In estuaries, a consideration, not typical

of streams and lakes, is one where the life cycle of anadromous and semi-anadromous fish species must

be considered as well as marine spawners that utilize estuaries as nursery areas. The RTAG may need to

consult with neighboring Regional RTAGs in regions where estuaries are shared. 

At present, EPA’s Office of Water is developing a policy to address effects from nutrient transport that

causes water quality problems at downstream estuarine sites including river systems that deliver nutrients

from far inland to coastal tidal systems. If downstream designated uses are not protected by a proposed

criterion, then the river or stream criterion must be modified accordingly.

Guidance for Interpreting and Applying Criteria

A critical step in the criteria development process is to assess how realistically criteria can be

implemented into standards that are accepted by the public. It should be realized that today’s designated

uses are not those that would be applicable in many estuaries at the turn of the century or in some cases

even several decades ago. Many estuaries have lost important fisheries that may not be easily recovered

if at all. For example, sturgeon are rare in many estuaries today when they were abundant decades ago in

several east coast estuaries. It is doubtful that the nutrient relationship for sturgeon growth and survival

is adequately known except for obvious factors such as hypoxia. The RTAG should make some

judgements about designated uses as exemplified by the sturgeon example that significantly improves

nutrient-based degraded water quality in terms of “fishable and swimmable” but maintains an important

degree of realism.

Do the Criteria Protect Designated Uses?

Section 303(c) of the CWA as amended (Public Law 92-500 [1972], 33 U.S.C. 1251, et seq.) requires all

States and authorized Tribes to establish designated uses for their waters. EPA’s interpretation of the

CWA requires that wherever attainable, standards should provide for protection and propagation of fish,

shellfish, and wildlife and provide for recreation in and on the water (section 101(a)). Note: this is the

secondary goal of the Act; the primary goal being the protection and restoration of the physical,

chemical, and biological integrity of the Nation’s waters, and zero discharge of pollution. Other uses

identified in the Act include industrial, agricultural, and public water supply. However, no waters may

be designated to be used as repositories for pollutants (see 40 CFR 131.10(a)). Each waterbody must

have criteria that protect and maintain the designated use of that water.

Also discussed below are general guidelines for developing criteria to protect selected designated uses. 

The values included here are not intended to represent proposed EPA or State estuarine and coastal

waters nutrient criteria. Rather, they are simply guidance and illustrate ranges of parameters associated

with the impairment of some designated uses in some States and Tribes. Criteria to protect these uses

should be developed on a site-specific basis when the individual nature of the estuary or coastal waters

require such specificity.
7-10 Nutrient Criteria—Estuarine and Coastal Waters

Outstanding National Resource Waters

Some estuarine and coastal waters of the State may require special criteria based on unique

characteristics of that waterbody. Such characteristics might include undisturbed or unique fjords or

subestuaries or stretches of coastline that are markedly different from other coastal waters in the State. 

Some areas may include threatened or endangered species that need to be protected. Such waters are the

very best of the reference set and are most in need of protection by rigid State and Tribal antidegradation

policies and procedures. 

Aquatic Life Uses

Aquatic life uses, including fisheries and shellfisheries, are heavily dependent on the initial high quality

condition of the resource. Species will change as a function of trophic state, and it may be difficult to

defend why one species is necessarily “better” than another. The use of reference areas and their

accompanying biota is one measure that can be used to predict the species that should be expected in a

region.

Fisheries

Developing criteria to protect a specific fishery may be somewhat difficult because in open estuarine and

coastal waters fish species shift with seasonal migrations and salinity changes. However, basic response

variables such as available DO and turbidity can be incorporated to protect all seasonal fish and

crustacean communities and resident molluscan populations. Consultation with fisheries managers, the

recreational public, and commercial fishermen should help resolve any issues of targeted species

management through nutrient abatement. 

Although our knowledge of the dynamics of change in the biota as a function of eutrophication requires

further development, there is sufficient evidence to conclude that eutrophication will bring species

changes. If an area has an existing aquatic life use, then that use must be maintained. (See 40 CFR

§131.12(a) (1).) Eutrophication will cause some species to change in relative abundance and cause

others to disappear; therefore, nutrient enrichment may be incompatible with the maintenance of a

specific biota. The ultimate extension of this concept is in the use classification of outstanding natural

resource waters. 

Recreation

Swimming/Primary Contact Recreation

Criteria to protect a contact recreation use may be associated with the occurrence (or appearance) of

certain phenomena that affect certain types of recreation. For example, in general, swimmers will not be

affected by the trophic state of the estuary, but resulting changes in transparency or change in species

may be important. Dense planktonic or macrophyte growth may inhibit swimming opportunities and help

promote seasonal densities of sea nettles or “jelly fish.” Excess nutrients feed not only nuisance algae

growth, but potentially health-endangering bacteria, especially when human or animal waste may be

involved such as when sewerage discharges are in the general area and may impact swimming when wind

and tide coordinate. This risk is not unusual for coastal ocean beaches where development promotes

sewerage expansion and offshore discharges.
Nutrient Criteria—Estuarine and Coastal Waters 7-11

Boating and Secondary Contact Recreation.

It might be expected that the transparency of the water or the presence of algal scums would not deter

boating, unless water skiing were involved. However, boating may be affected by the presence of dense

inshore beds of tall or floating macrophytes.

Restoration Goals

As described in the introduction to this chapter, designated uses may be protected but some nutrientbased ecological degradation already may have occurred. The public deserves to know what nutrient

conditions existed before anthropogenic nutrient enrichment initiated a shift from the natural nutrient

regime toward conditions of nutrient impairment within the limits of scientific knowledge or reasonable

scientific inference.

Sampling for Comparison to Criteria

Once criteria have been selected for each indicator variable (e.g., as a minimum, TN, TP, chl a, or

macroalgal biomass as AFDW and a measure of water clarity associated with chl a and, where

appropriate, the addition of dissolved oxygen), States and Tribes will want to develop implementation

procedures to assess the estuarine and coastal water with the criteria. Sampling to evaluate attainment

with criteria and adopted standards should be compatible with the procedures to establish the criteria in

the first place. If the criterion was developed for a particular season, then sampling should be compatible

with that season. In some cases, it is plausible that nutrient concentrations will not correlate predictably

with response variables because of estuarine hydrodynamics. In such cases, the useful relationship

should be directed toward nutrient loading. Many published estuarine nutrient relationships are based on

nutrient load, often normalized to estuarine surface area, not nutrient concentrations. In such cases, the

sampling of load relative to response variable should be scientifically based on the appropriate season

with consideration for appropriate time lags (see Chapter 2). In many cases, relationships will need to be

sought through application of a computer model (see Chapter 9). 

Questions will arise about the size of area, depth of sample, frequency, and duration of any exceedances. 

These are difficult questions but the RTAG must be prepared to address them. It is expected that

scientific judgment will be required in numerous cases that pushes the state of the science and in some

cases it may be necessary to make risk management decisions extend beyond the current state of science. 

Some illustrative examples are provided but should not necessarily be interpreted literally. If an area is

small and does not limit life cycle completion of important species (e.g., deepwater hypoxia that may

serve as a bottleneck to estuarine species migration), then some tolerance is accepted. However, if the

duration and magnitude of noncompliance of a criterion lasts long enough to affect the distribution and

abundance or recruitment of an important species or a key food web component at a designated level

(e.g., 15% reduction in the population of a harvestable size class is estimated based on the best available

judgment), then the criterion needs to be adjusted. In some cases, empirical or computer models will be

required to address many of the more complex relationships. 

The question arises, how many replicate samples are needed to obtain an acceptable precision of data in

order to detect differences between sites and changes over time? This depends on the nature of the
7-12 Nutrient Criteria—Estuarine and Coastal Waters

variability in the variable of interest. Several approaches are available. However, this question involves

both statistical and practical considerations (e.g., cost). General experience suggests that field water

quality sampling will often vary by 20%. With this “rule of thumb,” it may not be cost-effective to try to

achieve a lower percent difference. Eckblad (1991) provides some guidance on statistical considerations

in sampling power. The Kendall test with Sen slope estimate (Hirsch et al. 1982) allows the

determination of the number of replicate samples needed to detect a certain percent change in annual

means of a variable or a certain percent trend over a period such as 10 years (see Rivers and Streams

Nutrient Guidance Manual (U.S. EPA 2000b, Appendix A).

7.8 NUTRIENT CRITERIA INTERPRETATION PROCEDURES

However done, a State’s or Tribe’s nutrient criteria should include a procedural protocol to implement

the newly adopted nutrient criteria. The criteria and procedures should be reviewed by the RTAGs for

concurrence and are subject to further EPA review and approval if submitted as part of State or Tribal

standards.

The initial criteria variables include two causal variables (TN and TP) and two response variables (algal

biomas, e.g., chlorophyll a for phytoplankton and AFDW for macroalgae) and water clarity (e.g., Secchi

depth), and where hypoxia occurs, dissolved oxygen may be added as a third response variable. Failure

to meet either of the causal criteria should be sufficient to indicate a criteria “excursion,” and usually the

biological response, as measured by chlorophyll a and Secchi depth, will follow this nutrient trend. 

However, if the causal criteria are met but some combination of response criteria are not met, then there

should be some means of determining if the waters in question meet the nutrient criteria. Two suggested

approaches are described below.

Decisionmaking Protocol

One option is to establish a decisionmaking procedure equation all of the criteria. Such a rule might

state: “Both TN and TP causal nutrient criteria must be met, and a least three out of five response criteria

(e.g., water clarity, algal biomass as chlorophyll a or macroalgal biomass as AFDW, DO, seagrass or

SAV biomass, and phytoplankton species composition) must be met for three out of four sampling events

during the June through August survey period over 2 consecutive calendar years of sampling. No

sampling events may be less than 3 weeks apart [to avoid clustering sampling activities near a particular

flow condition or runoff event], and flow and tidal conditions must be recorded as well so that watershed

base flow and runoff events are evident and can be factored into the data assessment process.”

Multivariable Enrichment Index

The second option is to establish an index that accomplishes the same result by inserting the data into an

equation that relates the multiple variables in a nondimensional comprehensive score much the same way

an index of biotic integrity (Karr 1981) does. An example of an enrichment index approach is presented

in Table 7-1.
Nutrient Criteria—Estuarine and Coastal Waters 7-13

Table 7-1. Example of an enrichment index using the middle portion of a hypothetical estuary

Variable Criterion

Hypothetical estuary or salinity zone of estuary

Median measured value Enrichment Index (EI) score*

Causal variables

Total P (mg/L) #2.0 2.5 4

Total N (mg/L) #68 70 5

Primary response variables

Secchi depth (M) $1.0 0.6 3

Chlorophyll a (mg/L) #60 75 3

Secondary response variables

Dissolved oxygen (mg/L in

hypoliminion)

$6.0 3.5 4

Enrichment Index Value**: = 19

* Each of the eight variables receives an EI score. The scoring procedure is: 0 = meets criterion; 2 = fails to meet criterion by

10%; 3 = fails to meet criterion by 25%; 5 = fails to meet criterion by 50% or more.

** Enrichment Index Value is the sum of the EI scores. The maximum i.e. worst score achievable is 25.

If necessary, the scoring process can be weighted by seasons. Thus, different emphasis can be given to

the results of winter surveys as compared with summer surveys, and year-round work can be conducted if

necessary or desired. For example, greater weight perhaps by a factor of 2 could be given to the primary

response variables in winter for north temperate waters because these variables would normally be

expected to be improved at this time of year. Similarly, the criteria for TP and TN might both be

changed to lower concentration for winter because less runoff or fewer fertilizer applications are

expected in the watershed. In the example, the estuary or region thereof fails anyway because it failed

the criterion for either TP or TN (in fact it failed both). With a score of 19 out of a possible 25, it is also

a prime candidate for extensive remediation management.

Such enrichment index scores are not intended at this time to be surrogate nutrient criteria. They may,

however, serve as a “translator” to implement multiparameter criteria. However, like biological criteria

index scores such as the Index of Biotic Integrity, the enrichment index may be a useful assessment tool

relating several parameters. This helps the resource manager plan the distribution of effort and funds

over the entire estuarine or coastal resource base in one procedure.

Frequency and Duration

Frequency and duration are important concerns when evaluating any water with respect to meeting

criteria. This is a difficult process at this initial phase of the program because the data sources for

criteria development are presently so diverse. In general, however, the method of data gathering for

compliance should be as near as possible to that used to establish the criteria, especially keeping in mind
7-14 Nutrient Criteria—Estuarine and Coastal Waters

tidal phase and salinity. Once consistency is established, excursions from the criteria based on frequency

and duration can be evaluated whether based on a decision rule or a multivariable index.

Frequency of “excursion” from a criterion is a decision that can be best established by the State or Tribe

on the basis of their knowledge of the local water resources. An excursion that occurs less than 10

percent of the times when sampling is conducted (at regularly spaced or random intervals) may be

considered acceptable. Duration of the excursion may be stipulated as a set period of time (e.g., 2 weeks,

or as to not exist over more than two consecutive sampling intervals, whichever is the lesser period). The

State or Tribe in consultation with EPA will need to specifically define these terms as appropriate to the

region and should also determine the combination of these factors that constitutes an “excursion.”

7.9 CRITERIA MODIFICATIONS

Some situations may require site-specific criteria because of unique environmental conditions. In such

situations, the general criterion is a starting point and it must be modified to protect designated uses in a

unique situation. Such criteria can be adopted into State or Tribal water quality standards and reviewed

by EPA.

7.10 EPA, STATE, OR TRIBE RESPONSIBILITY UNDER THE CLEAN WATER ACT

The Clean Water Act as amended (Pub. L. 92-500 (1972), 33 U.S.C. 1251, et seq.) requires all States to

establish designated uses for their waters (Section 303(c)). Designated uses are set by the State. EPA’s

interpretation of the CWA requires that wherever attainable, standards should provide for the protection

and propagation of fish, shellfish, and wildlife and provide for recreation in and on the water (Section

101(a)). Other uses identified in the act include industrial, agricultural, and public water supply. 

However, no waters may be designated for use as repositories for pollutants (40 CFR 131.10 [a]). Each

waterbody must have criteria or measures of appropriate water quality that protect and maintain the

designated use of that water. It is recommended that the EPA nutrient guidance be followed. However,

States and Tribes may follow other guidance to adopt water quality criteria as long as the criteria are

based on scientifically sound methods and protect designated uses.

7.11 IMPLEMENTATION OF NUTRIENT CRITERIA INTO WATER 

QUALITY STANDARDS

Nutrient criteria adopted into water quality standards by States and Tribes are submitted to EPA for

review and approval (see Section 40 CFR 131). EPA reviews the criteria (40 CFR 131.5) for consistency

with the requirements of the CWA and 40 CFR 131.5, which requires that water quality criteria be

sufficient to protect the designated use (40 CFR 131.6 (c) and 40 CFR 131.11). The procedures for

State/Tribal review and revision of water quality standards, EPA review and approval of water quality

standards, and EPA promulgation of water quality standards (upon disapproval of State/Tribal water

quality standards) are found at 40 CFR 131.20-22. The Water Quality Standards Handbook (U.S. EPA

1994) provides guidance for implementation of these regulations.
Nutrient Criteria—Estuarine and Coastal Waters 8-1

Managing Point Source Pollution

Managing Nonpoint Source Pollution

Comprehensive Procedure for Nutrient

Management

Resources

CHAPTER 8

Using Nutrient Criteria To

Protect Water Quality

This chapter provides an introduction to the applications of nutrient criteria. Chapter 1 described the

ways in which nutrient criteria are used to (1) identify problems, (2) develop management plan, (3) assess

regulations, (4) evaluate projects, and (5) determine the status and trend of the water resource. In this

applications chapter, some of these are discussed in further detail. Section 8.1 addresses the management

of point source pollution, in the context of standards development, National Pollutant Discharge

Elimination System (NPDES) permits, and total maximum daily loads (TMDLs). Section 8.2 focuses

exclusively on nonpoint source management programs. Although some material is not directly related to

estuarine and coastal marine resources, it is included here because the coastal waters are the ultimate

recipients of all drainage, both coastal and inland, and the information may be useful to a manager

addressing various sources within a watershed. Coastal waters may, of course, be waters of the United

States (see 40 CFR 122.2) and may thus be subject to the requirements of the Clean Water Act. Section

8.3 sets out a comprehensive planning, application, and evaluation procedure for estuarine and coastal

marine nutrient quality management, and Section 8.4 lists publications on coastal/estuarine and

watershed management and protection.

8.1 MANAGING POINT SOURCE POLLUTION

The term "point source" means any discernible, confined, and discrete conveyance, including but not

limited to any pipe, ditch, channel, tunnel, conduit, well, discrete fissure, container, rolling stock,

concentrated animal feeding operation, or vessel or other floating craft, from which pollutants are or may

be discharged (CWA § 502(14)). This term does not include agricultural storm water discharges and

return flows from irrigated agriculture. This section describes some of the programs relevant to point

source discharges into rivers and streams.

The Clean Water Act and Water Quality Standards

The goals of the CWA are to achieve, wherever attainable, water quality that provides for protection and

propagation of fish, shellfish, and wildlife and recreation in and on the water. The CWA further

specifies that States adopt, and EPA approve, water quality standards consisting of designated uses,

criteria to project those uses, and an antidegradation policy (CWA section 303(c)). The criteria must be

based on a sound scientific rationale and must contain sufficient parameters or constituents to protect the

designated uses (40 CFR § 131.11(a)). For waters with multiple use designations, criteria must support

the most sensitive use (Id.). Finally, in designating uses and establishing water quality criteria, States

must ensure attainment of standards in downstream waters (40 CFR § 131.10(b)). With regard to

nutrient criteria, Section 304(a) of the CWA directs EPA to develop and publish criteria that reflect the

latest scientific knowledge of the effects of pollutants on biological community diversity, productivity,
8-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure 8-1. Components of water quality standards.

and stability, including information on the factors affecting rates of eutrophication for varying types of

receiving waters. In establishing water quality criteria, States should establish numeric values to protect

designated uses based on EPA’s Section 304(a) criteria guidance, modifications of the guidance

recommendations reflecting site-specific conditions, or criteria based on other scientifically defensible

methods (40 CFR § 131.11(b)(1)).

As illustrated in Figure 8-1, States adopt water quality standards for waters of the United States that

comprise designated uses, criteria to protect those uses, and an antidegradation policy to protect existing

water quality. Additionally, States develop implementation procedures to describe how the water quality

standards will be applied. Once water quality standards are adopted and approved, they become the basis

for legally enforceable NPDES permit limitations and a variety of assessment activities under the Clean

Water Act.

Protecting Designated Uses

It has been amply demonstrated that nutrients are a major contributor to use impairment in waters of the

United States. Because States are required to designate uses in consideration of the goals of the CWA

and adopt criteria that contain sufficient parameters to protect designated uses, and because it is EPA's

responsibility to make related recommendations, the Agency is developing and publishing Section 304(a)

criteria for nutrients that provide for protection and propagation of fish, shellfish, and wildlife and

recreation in and on the water.
Nutrient Criteria—Estuarine and Coastal Waters 8-3

EPA's Section 304(a) criteria for nutrients are issued on the basis of ecoregion and waterbody type. This

approach to nutrient criteria development provides a sound, scientifically defensible approach that

accounts for the characteristics of different types and locations of waterbodies. EPA's ecoregional

nutrient criteria are intended to represent enrichment conditions of surface waters minimally affected by

human development. These criteria may be developed and further refined on the basis of the five

elements described in this technical guidance manual.

Water quality criteria incorporating minimally affected (i.e., reference) conditions should provide for

protection and propagation of aquatic life and recreation and reflect conditions that will not adversely

affect the biological community. The parameters addressed in the Ecoregional Nutrient Criteria

Documents are total phosphorus, total nitrogen, chlorophyll a, and turbidity (e.g., Secchi depth for lakes;

turbidity for rivers and streams). These are the parameters that EPA considers important in nutrient

assessment because the first two (nitrogen and phosphorus) are the main causal agents of enrichment,

whereas the two response variables (chlorophyll a and turbidity) are indicators of system overenrichment

for most surface waters.

Maintaining Existing Water Quality

Antidegradation

State and Tribal water quality standards include an antidegradation policy and methods through which

the State or Tribe implements the policy. An antidegradation policy is required in State water quality

standards to protect existing water quality. At a minimum, States must maintain and protect the quality

of waters to support existing uses Antidegradation implementation procedures address the measures used

by States and Tribes to ensure that permits and control programs meet water quality standards and

antidegradation requirements. The water quality standards regulation sets out a three-tiered

antidegradation approach for the protection of water quality (40 CFR § 131.12). 

Tier 1

Maintains and protects existing uses and the water quality necessary to protect these uses (40 CFR

131.12[a][1]). An existing use can be established by demonstrating that fishing, swimming, or other uses

have actually occurred since November 28, 1975, or that the water quality is suitable to allow such uses

to occur, unless there are physical problems, such as substrate or flow, that prevent the use from being

attained (U.S. EPA 1994).

Tier 2

Protects the water quality in waters whose quality is better than that necessary to protect

"fishable/swimmable" uses (40 CFR 131.12[a][2]). The water quality standards regulation requires that

certain procedures be followed and certain showings be made (an "antidegradation review") before a

point source is authorized to lower water quality in high-quality waters. In no case may water quality for

a tier 2 waterbody be lowered to a level at which existing uses are impaired.
8-4 Nutrient Criteria—Estuarine and Coastal Waters

Tier 3

Outstanding national resource waters (ONRWs) are provided the highest level of protection under the

antidegradation policy. The policy provides for protection of water quality in high-quality waters that

constitute an ONRW by prohibiting the lowering of water quality. ONRWs are often regarded as highest

quality waters of the United States: That is clearly the thrust of 131.12(a)(3). However, ONRW

designation also offers special protection for waters of "exceptional ecological significance." These are

waterbodies that are important, unique, or sensitive ecologically, but whose water quality, as measured

by the traditional parameters such as dissolved oxygen or pH, may not be particularly high or whose

characteristics cannot be adequately described by these parameters (such as wetlands).

The regulation requires water quality to be maintained and protected in ONRWs. EPA interprets this

provision to mean no new or increased discharges to ONRWs and no new or increased discharge to

tributaries to ONRWs that would result in lower water quality in the ONRWs. The only exception to this

prohibition, as discussed in the preamble to the Water Quality Standards Regulation (48 FR 51402),

permits States to allow some limited activities that result in temporary and short-term changes in the

water quality of ONRW. Such activities must not permanently degrade water quality or result in water

quality lower than that necessary to protect the existing uses in the ONRW. It is difficult to give an exact

definition of "temporary" and "short-term" because of the variety of activities that might be considered. 

However, in rather broad terms, EPA's view of temporary is weeks and months, not years. The intent of

EPA's provision clearly is to limit water quality degradation to the shortest possible time. If a

construction activity is involved, for example, temporary is defined as the length of time necessary to

construct the facility and make it operational. During any period of time when, after opportunity for

public participation in the decision, the State allows temporary degradation, all practical means of

minimizing such degradation shall be implemented. Examples of situations in which flexibility is

appropriate can be found in the Water Quality Standards Handbook (U.S. EPA 1994).

General Policies

The water quality standards regulation allows States and Tribes to include implementation in their

standards policies and provisions, such as mixing zones, variances, and low-flow exemptions (40 CFR §

131.13). Such policies are subject to EPA review and approval. These policies and provisions should be

specified in the State’s or Tribe's water quality standards document. The rationale and supporting

documentation should be submitted to EPA for review during the water quality standards review and

approval process. 

Mixing Zones

States and Tribes may, at their discretion, allow mixing zones for dischargers. The water quality

standards should describe the methodology for determining the location, size, shape, outfall design, and

in-zone quality of mixing zones. Careful consideration should be given to the appropriateness of a

mixing zone where a substance discharged is bioaccumulative, persistent, carcinogenic, mutagenic, or

teratogenic.
Nutrient Criteria—Estuarine and Coastal Waters 8-5

Low-Flow Provisions

State and Tribal water quality standards should protect water quality for the designated and existing uses

in critical low-flow situations. States and Tribes may, however, designate a critical low-flow below

which numerical water quality criteria do not apply. When reviewing standards, States and Tribes should

review their low-flow provisions for conformance with EPA guidance. 

Water Quality Standards Variances

Variance procedures involve the same substantive and procedural requirements as removing a designated

use (see 40 CFR 131.10 (g)), but unlike use removal, variances are both discharger and pollutant specific,

are time-limited, and do not forego the currently designated use.

A variance should be used instead of removal of a use where the State believes the standard can

ultimately be attained. By maintaining the standard rather than changing it, the State will assure that

further progress is made in improving water quality and attaining the standard. With a variance, NPDES

permits may be written such that reasonable progress is made toward attaining the standards without

violating section 402(a)(1) of the Act, which requires that NPDES permits must meet the applicable

water quality standards.

State variance procedures, as part of State water quality standards, must be consistent with the

substantive requirements of 40 CFR 131. EPA has approved State-adopted variances in the past and will

continue to do so if: 

• Each individual variance is included as part of the water quality standard

• The State demonstrates that meeting the standard is unattainable based on one or more of the grounds

outlined in 40 CFR 131.10(g) for removing a designated use

• The justification submitted by the State includes documentation that treatment more advanced

than that required by sections 303(c)(2)(A) and (B) has been carefully considered, and that

alternative effluent control strategies have been evaluated

• The more stringent State criterion is maintained and is binding upon all other dischargers on the

stream or stream segment

• The discharger who is given a variance for one particular constituent is required to meet the

applicable criteria for other constituents

• The variance is granted for a specific period of time and must be rejustified upon expiration but at

least every 3 years (Note: the 3-year limit is derived from the triennial review requirements of section

303(c) of the Act.)
8-6 Nutrient Criteria—Estuarine and Coastal Waters

• The discharger either must meet the standard upon the expiration of this time period or must make a

new demonstration of "unattainability"

• Reasonable progress is being made toward meeting the standards

• The variance was subjected to public notice, opportunity for comment, and public hearing. (See

section 303(c)(1) and 40 CFR 131.20.) The public notice should contain a clear description of the

impact of the variance on achieving water quality standards in the affected stream segment.

Providing Flexibility in Implementation

Abundant flexibility is built into the criteria-setting process and water quality standards regulations to

allow States to (1) develop their own criteria to protect specific uses or reflect local conditions, (2) use

different techniques to develop criteria as long as they are protective and scientifically defensible, and (3)

conduct use attainability studies and refine their use designations.

States also have the flexibility to adopt numeric criteria to protect designated uses or adopt methods and

procedures that "translate" narrative criteria into numeric values. Narrative criteria statements, often

referred to as "general criteria" in States' standards regulations, usually take the form of a description of

desired water quality condition or a preclusion of certain types of pollution or undesirable conditions

(i.e., the "free from" provisions). Narrative criteria are considered critical backstops for designated use

protection and are a powerful means of achieving desired water quality if they are interpreted in a clear

and consistent manner. In water quality standards parlance, a "translator" is a process, methodology, or

guide that States or Tribes use to quantitatively interpret narrative criteria statements. Translators may

consist of biological assessment methods (e.g., field measures of the biological community), biological

monitoring methods (e.g., laboratory toxicity tests), models or formulas that use input of site-specific

information/data, or other scientifically defensible methods. Translators are particularly useful in

describing water quality conditions that require a greater degree of sophistication to assess than typically

can be expressed by numerical criteria that apply broadly to all waters with a given use designation. The

translator may be either directly incorporated into State or Tribal water quality standards or incorporated

by reference. In either case, specific limits or values for a measurable pollutant derived using a translator

that interpret a narrative criterion statement should be attached to the State or Tribal regulations to ensure

public review, as would be required of any site-specific numerical criterion.

States have the flexibility under current law to adopt appropriate nutrient criteria. If a State determines

that its water quality criteria cannot be met, that State may consider refining use designation, adopting

site-specific criteria, or issuing a variance to ensure that the appropriate uses and criteria to protect the

uses are established. For example, if a regulated source faces expensive treatment to comply with a new

or revised requirement, the State or Tribe can authorize a variance, based on a justification using one of

the six factors at 131.10(g), to allow time for the discharge to come into compliance with a permit limit

for the criteria and/or weigh options on treatment technologies, use reclassification, or site-specific

development criteria as appropriate.
Nutrient Criteria—Estuarine and Coastal Waters 8-7

Water quality criteria guidance published by EPA under Section 304(a) of the CWA, such as for

nutrients, serve as primary sources of information to States and Tribes as they develop numeric criteria as

part of their water quality standards. Under the CWA and EPA's implementing regulations, States and

Tribes also may use other information, including local water quality conditions, as they develop

standards. Typically, EPA uses its own water quality criteria guidance as the principal basis for

proposing and promulgating a replacement water quality standard when a State or Tribe fails to adopt an

acceptable standard. In doing so, EPA commits to a process that includes public review and comment. 

EPA will solicit information from the public to determine if any such proposed Federal nutrient criteria

for State waters are sufficiently protective of uses. This public process will help ensure that any

promulgated Federal water quality standards are appropriately protective.

NPDES Permits

The CWA requires wastewater dischargers to have a permit establishing effluent limits on pollutant

discharges. The regulations at 40 CFR 122.41 et seq. require these permits to specify monitoring and

reporting requirements. More than 200,000 sources are regulated by the NPDES permits nationwide. 

These permits regulate household and industrial wastes that are collected in sewers and treated at

municipal wastewater treatment plants. Permits also regulate industrial point sources and concentrated

animal feeding operations that discharge into other wastewater collection systems or that have the

potential to discharge directly into receiving waters. Permits regulate discharges with the goals of

protecting public health and aquatic life and ensuring that every facility treats wastewater. Typical

pollutants regulated by NPDES are "conventional pollutants" such as fecal coliforms or oil and grease

from the sanitary wastes of households, businesses, and industries; and "toxic pollutants" including

pesticides, solvents, polychlorinated biphenyls (PCBs), dioxins, and heavy metals that are particularly

harmful to animal or plant life. "Nonconventional pollutants" are any additional substances that are not

conventional or toxic that may require regulation, including nutrients such as N and P. 

[Source: http://www.epa.gov/owm/gen2.htm]

Discharge monitoring data for pollutants limited and/or monitored pursuant to NPDES permits issued by

States, Tribes, or EPA are required to be stored in the central EPA Permit Compliance System (PCS). 

The assessment of point source loadings is not a simple process of assessing PCS data, even though PCS

is an important data source. The PCS database does not provide complete information for important N

sources. Most PCS N data are generated by water quality-based permit limitations on ammonia, often

applied in discharges to smaller streams. Few data exist in PCS on other forms of N, or TN; and data for

TP is not frequently found in PCS. This situation exists largely because most permits do not include

limits and/or monitoring requirements for N or P. The lack of nutrient limits and/or monitoring

requirements in permits is due to a general lack of State water quality standards for these parameters. 

[Source: http://www.epa.gov/msbasin/protocol.html]

The NPDES Storm Water Permitting Program

Storm water runoff is one of the remaining causes of contaminated lakes, streams, rivers, and estuaries

throughout the country. Pollution in storm water runoff is responsible for closing beaches and shellfish

harvesting areas, contaminating fish, and reducing populations of water plants and other aquatic life. 
8-8 Nutrient Criteria—Estuarine and Coastal Waters

High flows of storm water runoff cause flooding, property damage, erosion, and heavy siltation. The

1987 Congressional Amendments to the Clean Water Act required EPA to control pollution from storm

water discharges. In 1990, EPA promulgated Phase I regulations to control storm water discharges from

municipal separate storm sewer systems (MS4s) serving populations of more than 100,000, construction

activities disturbing more than 5 acres, and industrial facilities through issuance of NPDES storm water

permits. EPA promulgated Phase II of the program in 1999 to control storm water discharges from MF4s

less than 100,000 and small construction sites between 1 and 5 acres in size. The Phase II regulations

also expanded the exemption for industrial facilities that do not have exposure of industrial activities and

materials to storm water.

Construction Permits

The 1987 Congressional Amendments to the CWA required EPA to control pollution from storm water

discharges. EPA issued a general NPDES permit for construction sites disturbing 5 or more acres in

1992. General permits provide EPA with an effective mechanism to regulate discharges from tens of

thousands of construction sites, thus protecting and improving surface water quality across the Nation. 

Several general permits for construction activity have been issued/reissued since that first permit in 1992. 

EPA Regions 1, 2, 3, 7, 8, 9, and 10 have a general permit that authorizes the discharge of storm water

associated with construction activity disturbing 5 or more acres and smaller sources as designated by the

Agency on a case-by-case basis. This multiregional permit is known as the "Construction General

Permit" (CGP). 

Region 4 has issued a separate CGP for the State of Florida and Indian Country lands in Florida,

Mississippi, Alabama, and North Carolina. Region 6 has also issued its own CGP for the States of Texas

and New Mexico; Indian Country lands in Texas, New Mexico, Oklahoma, and Louisiana; and

construction activity at oil, gas, and pipeline facilities in Oklahoma.

As used in these construction permits, the term "storm water associated with construction activity" refers

to category (x) of the definition of "discharge of storm water associated with industrial activity," which

includes construction sites and common plans of development or sale that disturb 5 or more acres (see 40

CFR 122.26 [b][14]). The CGP applies only to areas for which EPA is the permitting authority (certain

States, Federal facilities, and Indian Lands). The majority of the country (i.e., 44 States and the Virgin

Islands) has been granted authority for permitting storm water discharges and as such, each of these

States is required to develop permits to control discharges from construction activities. In response to the

Phase II regulations, permit applications from construction activities between 1 and 5 acres is required by

March 2003.

Combined Sewer Overflows (CSOs)

Combined sewer overflows, or CSOs, can be a significant water pollution and public health threat in

urban areas. EPA's 1994 CSO Control Policy is a comprehensive national strategy to ensure that cities,

NPDES authorities, water quality standards authorities, and the public engage in a comprehensive and

coordinated planning effort to implement cost-effective CSO controls that meet the objectives and

requirements of the Clean Water Act.
Nutrient Criteria—Estuarine and Coastal Waters 8-9

During dry weather, combined sewer systems transport wastewater directly to sewage treatment plants. 

In periods of rainfall or snowmelt, however, the wastewater volume in a combined sewer system can

exceed the capacity of the collection system or treatment plant. When this happens, combined sewer

systems overflow and discharge untreated wastewater directly to streams, rivers, lakes, or estuaries.

It provides guidance to municipalities and State and Federal permitting authorities on meeting pollution

control goals of the CWA in a flexible, cost-effective manner. Information on EPA's CSO Control

Policy, accompanying guidance documents, and other elements of the national CSO control program can

be found on the following website: http://www.epa.gov/npdes.htm.

Storm Water Planning

The Watershed Management Institute, Inc., recently published a new manual entitled Operation,

Maintenance, and Management of Stormwater Management Systems (1998). This manual presents a

comprehensive review of the technical, educational, and institutional elements needed to ensure that

storm water management systems are designed, built, maintained, and operated properly during and after

their construction. The manual was developed in cooperation with the EPA Office of Water to assist

individuals responsible for designing, building, maintaining, or operating storm water management

systems. It will also be helpful to individuals responsible for implementing urban storm water

management programs.

The manual includes fact sheets on 13 common storm water treatment best management practices

(BMPs) that summarize operation, maintenance, and management needs and obligations, along with

construction recommendations. Other chapters review planning and design considerations, programmatic

and regulatory aspects, considerations for facility owners, construction inspection, inspection and

maintenance after construction, costs and financing, and disposal of storm water sediments. Forms for

inspecting BMPs during construction and determining maintenance needs afterwards are included in the

manual and in a separate supplement. 

[Source: http://www.epa.gov/owowwtr1/NPS/wmi/index.html]

[Additional information: http://www.epa.gov/owowwtr1/NPS/ordinance/osm6.htm and

http://www.epa.gov/owowwtr1/info/NewsNotes/issue05/nps05sto.html]

Total Maximum Daily Load (TMDL)

States, territories, and authorized Tribes establish section 303(d) lists of impaired waters based on

information contained in their 305(b) reports as well as other relevant and available water quality data. 

The section 303(d) list is a prioritized list of waters not meeting water quality standards. States are

required to submit lists biennially (40 CFR § 130.7(d)). States must develop TMDLs for waters and

pollutants on their section 303(d) lists (CWA § 303(d)(1)(C)). State section 303(d) lists and TMDLs

must be submitted to EPA for approval or disapproval. If EPA disapproves a list or TMDL submission,

EPA must identify waters for the list or establish the TMDL itself (CWA § 303(d)(2)).

A TMDL is a written, quantitative plan and analysis for attaining and maintaining water quality standards

in all seasons for a specific waterbody and pollutant. Specifically, a TMDL is the sum of the allowable
8-10 Nutrient Criteria—Estuarine and Coastal Waters

loads of a pollutant from all contributing point, nonpoint, and background sources (40 CFR § 130.2(i)). 

TMDLs may be established on a coordinated basis for a group of waterbodies in a watershed. TMDLs

identify the loading capacity of the water, wasteload allocations (for point sources), load allocations (for

nonpoint sources), and a margin of safety (U.S. EPA 1999) (40 CFR § 130.2), and are calculated at levels

necessary to achieve applicable water quality standards (CWA § 303(d)(1)(C)).

A waste load allocation (WLA) is the proportion of a receiving water's TMDL that is allocated to point

sources of pollution. Water quality models are often utilized by regulatory agencies in conducting an

assessment to determine a WLA. Models establish a quantitative relationship between a waste load and

its impact on water quality. WLAs are used by permit writers to establish Water Quality Based Effluent

Limits (WQBELs).

[Source: http://www.epa.gov:80/owmitnet/permits/pwcourse/chapt_06.pdf]

Both the 1996 and 1998 section 303(d) lists, as well as more recent 305(b) reports, reflect similar

patterns: sediments, nutrients, and pathogens are the top three causes of waterbody impairment.

[Source: http://www.epa.gov/owowwtr1/tmdl/faq.html]

Continuing Planning Process (CPP)

Each State is required to establish and maintain a continuing planning process (CPP) as described in

section 303(e) of the CWA. A State's CPP contains, among other items, a description of the process that

the State uses to identify waters needing water quality-based controls, the process for developing

TMDLs, and a description of the process used to receive public review of each TMDL (40 CFR § 130.5

& 130.7(a) & (c)). Descriptions may be as detailed as the Regional office and the State determine is

necessary to describe each step of the TMDL development process. This process may be included as part

of the EPA/State Agreement for TMDL development.

[Source: http://www.epa.gov/owowwtr1/tmdl/decisions/dec4.html]

Look to the Future ... Pollutant Trading

Point and nonpoint source pollutant trading involves financing reductions in nonpoint source pollution in

lieu of undertaking more expensive point source pollution reduction efforts. A trading program is

intended to produce cost savings for point source dischargers while improving water quality. In order for

a trading program to be viable, there should be a waterbody identifiable as a watershed or segment, as

well as a measurable combination of point sources and controllable nonpoint sources. In addition, point

source dischargers can be expected to trade for nonpoint source reductions if they perceive this as an

alternative to upgrade facility treatment capabilities. In addition, there should be significant load

reductions for which the cost per pound reduced for nonpoint source controls is lower than the cost for

upgrading point source controls.

Such a program allows the private sector to allocate its resources to reduce pollutants in the most

cost-effective manner, and it encourages the development of a watershed-wide or basin-wide

approach to water quality protection. A pollutant trading program also requires cooperation between

agencies and requires a system to arrive at trading ratios between point and nonpoint source controls.
Nutrient Criteria—Estuarine and Coastal Waters 8-11

For example, in a North Carolina watershed, the Tar-Pamlico Basin Association (a coalition of point

source dischargers) and State and regional environmental groups have proposed a two-phased nutrient

management strategy that incorporates point and nonpoint source pollutant trading. The plan requires

association members to finance nonpoint source reduction activities in the basin if their nutrient

discharges exceed a base allowance.

[Source: http://www.epa.gov/OWOW/NPS/MMGI/funding.html#9]

8.2 MANAGING NONPOINT SOURCE POLLUTION

During the first 15 years of the national program to abate and control water pollution, EPA and the States

focused most of their water pollution control activities on traditional "point sources," such as discharges

through pipes from sewage treatment plants and industrial facilities. These point sources have been

regulated by EPA and the States through the NPDES permit program established by section 402 of the

CWA. Discharges of dredged and fill materials into wetlands have also been regulated by the U.S. Army

Corps of Engineers and EPA under section 404 of the Clean Water Act.

The Nation has greatly reduced pollutant loads from point source discharges and has made considerable

progress in restoring and maintaining water quality as a result of the above activities. However, the gains

in controlling point sources have not solved all of the Nation's water quality problems. Recent studies

and surveys by EPA and by State/Tribal water quality agencies indicate that the majority of the

remaining water quality impairments in our Nation's rivers, streams, lakes, estuaries, coastal waters, and

wetlands result from nonpoint source pollution and other nontraditional sources, such as urban storm

water discharges and combined sewer overflows.

Nonpoint source pollution generally results from land runoff, precipitation, atmospheric deposition,

drainage, seepage, or hydrologic modification. Technically, the term "nonpoint source" is defined to

mean any source of water pollution that does not meet the legal definition of "point source" in section

502(14) of the CWA, defined in the preceding section. Although diffuse runoff is generally treated as

nonpoint source pollution, runoff that enters and is discharged from conveyances such as those described

above is treated as a point source discharge and hence is subject to the permit requirements of the Clean

Water Act. In contrast, nonpoint sources are not subject to Federal permit requirements.

The pollution of waters by nonpoint sources is caused by rainfall or snowmelt moving over and through

the ground. As the runoff moves, it picks up and carries away natural pollutants and pollutants resulting

from human activity, finally depositing them into lakes, rivers, wetlands, coastal waters, and

groundwaters. Nonpoint source pollution can also be caused by atmospheric deposition of pollutants

onto waterbodies. Furthermore, hydrologic modification is a form of nonpoint source pollution that often

adversely affects the biological and physical integrity of surface waters. A more detailed discussion of

the range of nonpoint sources and their effects on water quality and riparian habitats is provided in

subsequent chapters of this guidance. A summary of State laws related to nonpoint source pollution can

be found in the Almanac of Enforceable State Laws to Control Nonpoint Source Water Pollution (ELI

1988). This report can be accessed on the Internet at http://www.eli.org/bookstore/research.htm.
8-12 Nutrient Criteria—Estuarine and Coastal Waters

Nonpoint Sources of Nutrients

Guidance Specifying Management Measures for Sources of Nonpoint Pollution in Coastal Waters (U.S.

EPA 1993a) was developed by EPA for the planning and implementation of Coastal Nonpoint Pollution

Programs. The guidance focuses on controlling five major categories of nonpoint sources that impair or

threaten waters nationally. Management measures are specified for (1) agricultural runoff, (2) urban

runoff (including developing and developed areas), (3) silvicultural (forestry) runoff, (4) marinas and

recreational boating, and (5) hydromodification (e.g., channelization and channel modification, dams,

and streambank and shoreline erosion). EPA guidance also includes management measures for wetlands,

riparian areas, and vegetated treatment systems that apply generally to various categories of sources of

nonpoint pollution. Management measures are defined in the Coastal Zone Act Reauthorization

Amendments of 1990 (CZARA) as economically achievable measures to control the addition of

pollutants to waters, which reflect the greatest degree of pollutant reduction achievable through the

application of the best available nonpoint pollution control practices, technologies, processes, siting

criteria, operating methods, or other alternatives.

The following section outlines some of the management measures specified in the CZARA guidance for

the various types of nonpoint sources. These measures should be considered when implementing

programs targeting nutrient releases into waters of the United States.

Agricultural Runoff

• Erosion and sediment control

• Control of facility wastewater and runoff from confined animal facilities

• Nutrient management planning on cropland

• Grazing management systems

• Irrigation water management

Urban Runoff

• Control of runoff and erosion from existing and developing areas

• Construction site runoff and erosion control

• Construction site chemical control (includes fertilizers)

• Proper design, location, installation, operation, and maintenance of on-site disposal systems

• Pollution prevention education (e.g., household chemicals, lawn and garden activities, golf courses,

pet waste, on-site disposal systems, etc.)

• Planning, siting, and developing roads, highways, and bridges (including runoff management)

Silvicultural Runoff

• Streamside management

• Road construction and management

• Forest chemical management (includes fertilizers)

• Revegetation

• Preharvest planning, harvesting management
Nutrient Criteria—Estuarine and Coastal Waters 8-13

Marinas and Recreational Boating

• Siting and design

• Operation and maintenance

• Storm water runoff management

• Sewage facility management

• Fish waste management

• Pollution prevention education (e.g., proper boat cleaning, fish waste disposal, and sewage pumpout

procedures)

Hydromodification (i.e., channelization, channel modification, dams)

• Minimize changes in sediment supply and pollutant delivery rates through careful planning and

design

• Erosion and sediment control

• Chemical and pollutant control (includes nutrients)

• Stabilization and protection of eroding streambanks or shorelines

Wetlands, Riparian Areas, Vegetated Treatment Systems

• Protect the NPS abatement and other functions of wetlands and riparian areas through vegetative

composition and cover, hydrology of surface and groundwater, geochemistry of the substrate, and

species composition

• Promote restoration of preexisting function of damaged and destroyed wetlands and riparian systems

• Promote the use of engineered vegetated treatment systems if they can serve a NPS pollution

abatement function

Efforts To Control Nonpoint Source Pollution

Efforts to control nonpoint source pollution include nonpoint source management programs, the National

Estuary Program, atmospheric deposition, coastal nonpoint pollution control programs, and Farm Bill

conservation provisions. These efforts are described below.

Nonpoint Source Management Programs

In 1987, in view of the progress achieved in controlling point sources and the growing national

awareness of the increasingly dominant influence of nonpoint source pollution on water quality,

Congress amended the Clean Water Act to focus greater national efforts on nonpoint sources. In the

Water Quality Act of 1987, Congress amended section 101, “Declaration of Goals and Policy,” to add the

following fundamental principle:

It is the national policy that programs for the control of nonpoint sources of

pollution be developed and implemented in an expeditious manner so as to enable

the goals of this Act to be met through the control of both point and nonpoint sources

of pollution.

More importantly, Congress enacted section 319 of the CWA, which established a national program to

control nonpoint sources of water pollution. Under section 319, States address nonpoint pollution by
8-14 Nutrient Criteria—Estuarine and Coastal Waters

assessing nonpoint source pollution problems and causes within the State, adopting management

programs to control the nonpoint source pollution, and implementing the management programs. 

Although not required, many States have incorporated the management measures specified in the 1993

CZARA guidance into their State Nonpoint Source Management Programs. 

Section 319 also authorizes EPA to issue grants to States to assist them in implementing those

management programs or portions of management programs that have been approved by EPA. As of FY

2000, more than $1 billion in grants have been given to States, Territories, and Tribes for the

implementation of nonpoint source pollution control programs.

For additional information on the Nonpoint Source Management Program and distribution of Section 319

grants in your State, contact your State's designated nonpoint source agency. For many States, the

nonpoint source agency is the State Water Quality Agency. However, in several instances, other

agencies or departments are given nonpoint source responsibility (see Table 8-1).

National Estuary Program 

EPA also administers the National Estuary Program under section 320 of the CWA. This program

focuses on point and nonpoint pollution in geographically targeted, high-priority estuarine waters. Under

this program, EPA assists State, regional, and local governments in developing comprehensive

conservation and management plans that recommend priority corrective actions to restore estuarine water

quality, fish populations, and other designated uses of the waters. For additional information, contact

your local estuary program. The following estuaries are currently enrolled in the program:

Table 8-1. States for which the nonpoint source agency is not the water quality agency

State State Nonpoint Source Agency

Arkansas State Department of Soil and Water Conservation

Delaware State Department of Soil and Water Conservation

Oklahoma State Department of Soil and Water Conservation

Tennessee State Department of Agriculture

Texas Department of Soil and Water Conservation (for agriculture)

Texas Water Quality Board (all other nonpoint sources)

Vermont State Department of Agriculture

Virginia State Department of Soil and Water Conservation
Nutrient Criteria—Estuarine and Coastal Waters 8-15

• Albemarle-Pamlico Sounds, NC

• Barataria-Terrebonne Estuarine Complex, LA

• Barnegat Bay, NJ

• Buzzards Bay, MA

• Casco Bay, ME

• Charlotte Harbor, FL

• (Lower) Columbia River Estuary, OR and WA

• Corpus Christi Bay, TX

• Delaware Estuary, DE, NJ, and PA

• Delaware Inland Bays, DE

• Galveston Bay, TX

• Indian River Lagoon, FL

• Long Island Sound, NY and CT

• Maryland Coastal Bays, MD

• Massachusetts Bays, MA

• Mobile Bay, AL

• Morro Bay, CA

• Narragansett Bay, RI

• New Hampshire Estuaries, NH

• New York-New Jersey Harbor, NY and NJ

• Peconic Bay, NY

• Puget Sound, WA

• San Francisco Estuary, CA

• San Juan Bay, PR

• Santa Monica Bay, CA

• Sarasota Bay, FL

• Tampa Bay, FL

• Tillamook Bay, OR

Atmospheric Deposition

Even though runoff from agricultural and urban areas may be the largest source of nonpoint pollution,

growing evidence suggests that atmospheric deposition may have a significant influence on nutrient

enrichment, particularly from nitrogen (Jaworski et al. 1997). Gases released through fossil fuel

combustion and agricultural practices are two major sources of atmospheric N that may be deposited in

waterbodies (Carpenter et al. 1998). Nitrogen and nitrogen compounds formed in the atmosphere return

to the earth as acid rain or snow, gas, or dry particles (http://www.epa.gov/acidrain/effects/envben.html). 

EPA has several programs that address the issue of atmospheric deposition, including the National

Ambient Air Quality Standards, the Atmospheric Deposition Initiative, and the Great Waters Program. 

National Ambient Air Quality Standards

The Clean Air Act provides the principal framework for national, State, and local efforts to protect air

quality. Under the Clean Air Act, national ambient air quality standards (NAAQS) for pollutants that are

considered harmful to people and the environment are established.

The Clean Air Act established two types of national air quality standards. Primary standards set limits to

protect public health, including the health of "sensitive" populations such as asthmatics, children, and the

elderly. Secondary standards set limits to protect public welfare, including protection against decreased

visibility and damage to animals, crops, vegetation, and buildings (http://www.epa.gov/airs/criteria.html).

Atmospheric Deposition Initiative

In 1995, EPA's Office of Water established an "Air Deposition Initiative" to work with the EPA Office of

Air and Radiation to identify and characterize air deposition problems with greater certainty and examine

solutions to address them. The Air and Water Programs are cooperating to assess the atmospheric

deposition problem, conduct scientific research, provide innovative solutions to link Clean Air Act and

Clean Water Act tools to reduce the of these pollutants, and communicate the findings to the public. To

date, most efforts have focused on better understanding of the links between nitrogen and mercury
8-16 Nutrient Criteria—Estuarine and Coastal Waters

emissions and harmful effects on water quality and the environment. Significant work has also been done

towards quantifying the benefits to water quality of reducing air emissions and developing sensible, costeffective approaches to reducing the emissions and their ecosystem and health effects

(http://www.epa.gov/owowwtr1/oceans/airdep/index.html).

Great Waters Program

On November 15, 1990, in response to mounting evidence that air pollution contributes to water

pollution, Congress amended the Clean Air Act and included provisions that established research and

reporting requirements related to the deposition of hazardous air pollutants to the "Great Waters." The

waterbodies designated by these provisions are the Great Lakes, Lake Champlain, and Chesapeake Bay. 

As part of the Great Waters Program, Congress requires EPA, in cooperation with the National Oceanic

and Atmospheric Administration, to monitor hazardous pollutants by establishing sampling networks, 

investigate the deposition of these pollutants, improve monitoring methods, monitor for hazardous

pollutants in fish and wildlife, determine the contribution of air pollution to total pollution in the Great

Waters, evaluate any adverse effects to public health and the environment, determine sources of

pollution, and provide a report to Congress every 2 years. These reports provide an information base that

can be used to establish whether air pollution is a significant contributor to water quality problems of the

Great Waters, determine whether there are significant adverse effects to humans or the environment,

evaluate the effectiveness of existing regulatory programs in addressing these problems, and assess

whether additional regulatory actions are needed to reduce atmospheric deposition to the Great Waters. 

For more detail, the Great Waters biennial reports to Congress discuss current scientific understanding of

atmospheric deposition (http://www.epa.gov/airprogm/oar/oaqps/gr8water/xbrochure/program.html).

Coastal Nonpoint Pollution Control Programs

In November 1990, Congress enacted CZARA. These amendments were intended to address several

concerns, a major one being the impact of nonpoint source pollution on coastal waters.

To address more specifically the impacts of nonpoint source pollution on coastal water quality, Congress

enacted section 6217, "Protecting Coastal Waters," which was codified as 16 USC-1455b. This section

provides that each State with an approved coastal zone management program must develop and submit a

Coastal Nonpoint Pollution Control Program for EPA and National Oceanic and Atmospheric

Administration (NOAA) approval. The purpose of the program "shall be to develop and implement

management measures for nonpoint source pollution to restore and protect coastal waters, working in

close conjunction with other State and local authorities."

States with Coast Nonpoint Pollution Control Programs are required to include measures in their

programs that are "in conformity" with the 1993 CZARA guidance, as discussed previously. A listing of

States with Coastal Nonpoint Pollution Control Programs is presented in Table 8-2. For additional

information on the programs in these States, contact the State water quality agency.
Nutrient Criteria—Estuarine and Coastal Waters 8-17

Table 8-2. States and Territories with coastal nonpoint pollution control programs

Alabama Maine Oregon

Alaska Maryland Pennsylvania 

American Samoa Massachusetts Puerto Rico 

California Michigan Rhode Island

Connecticut Mississippi South Carolina

Delaware New Hampshire Virgin Islands

Florida New Jersey Virginia 

Guam New York Washington

Hawaii North Carolina Wisconsin 

Louisiana Northern Mariana Islands

Farm Bill Conservation Provisions

Technical and financial assistance for landowners seeking to preserve soil and other natural resources is

authorized by the Federal Government under provisions of the Food Security Act (Farm Bill). Provisions

of the 1996 Farm Bill relating directly to installation and maintenance of BMPs are summarized in the

following sections. Contact your Natural Resources Conservation Service (NRCS) State

Conservationist's office for State-specific information.

Environmental Conservation Acreage Reserve Program (ECARP)

ECARP is an umbrella program established by the 1996 Farm Bill and contains the Conservation Reserve

Program (CRP), Wetlands Reserve Program (WRP), and Environmental Quality Incentives Program

(EQIP). It authorizes the Secretary of Agriculture to designate watersheds, multi-State areas, or regions

of special environmental sensitivity as conservation priority areas eligible for enhanced Federal

assistance. Assistance in priority areas is to be used to help agricultural producers comply with NPS

pollution requirements of the CWA and other State or Federal environmental laws. The ECARP is

authorized through 2002.

Conservation Reserve Program (CRP)

First authorized by the Food Security Act of 1985 (Farm Bill), this voluntary program offers annual

rental payments, incentive payments, and cost-share assistance for establishing long-term,

resource-conserving cover crops on highly erodible land. CRP contracts are issued for a duration of 10 to

15 years for up to 36.4 million acres of cropland and marginal pasture. Land can be accepted into the

CRP through a competitive bidding process through which all offers are ranked using an environmental

benefits index, or through continuous signup for eligible lands where certain special conservation

practices will be implemented.

The Conservation Reserve Enhancement Program (CREP) is a new initiative of CRP authorized under

the 1996 Federal Agricultural Improvement and Reform Act. CREP is a joint, State-Federal program

designed to meet specific conservation objectives. CREP targets State and Federal funds to achieve
8-18 Nutrient Criteria—Estuarine and Coastal Waters

shared environmental goals of national and State significance. The program uses financial incentives to

encourage farmers and ranchers to voluntarily protect soil, water, and wildlife resources.

Wetlands Reserve Program (WRP)

The WRP is a voluntary program to restore and protect wetlands and associated lands. Participants may

sell a permanent or 30-year conservation easement or enter into a 10-year cost-share agreement with

USDA to restore and protect wetlands. The landowner voluntarily limits future use of the land, yet

retains private ownership. The NRCS provides technical assistance in developing a plan for restoration

and maintenance of the land. The landowner retains the right to control access to the land and may lease

the land for hunting, fishing, and other undeveloped recreational activities.

Environmental Quality Incentives Program (EQIP)

The EQIP was established by the 1996 Farm Bill to provide a voluntary conservation program for

farmers and ranchers who face serious threats to soil, water, and related natural resources. EQIP offers

financial, technical, and educational help to install or implement structural, vegetative, and management

practices designed to conserve soil and other natural resources. Current priorities for these funds dictate

that one-half of the available monies be directed to livestock-related concerns. Cost-sharing may pay up

to 75% of the costs for certain conservation practices. Incentive payments may be made to encourage

producers to perform land management practices such as nutrient management, manure management,

integrated pest management, irrigation water management, and wildlife habitat management.

Wildlife Habitat Incentives Program (WHIP)

This program is designed for parties interested in developing and improving wildlife habitat on private

lands. Plans are developed in consultation with NRCS and the local Conservation District. USDA will

provide technical assistance and share up to 75% of the cost of implementing the wildlife conservation

practices. Participants generally must sign a 5- to 10-year contract with USDA that requires they

maintain the improvement practices.

Forestry Incentives Program (FIP)

Originally authorized in 1978, the FIP allows cost sharing of up to 65% (up to a maximum of $10,000 per

person per year) for tree planting, timber stand improvement, and related practices on nonindustrial

private forest land. The FIP is administered by NRCS and the U.S. Forest Service. Cost share funds are

restricted to individuals who own no more than 1,000 acres of eligible forest land.

Conservation of Private Grazing Land

This program was authorized by the 1996 Farm Bill for providing technical and educational assistance to

owners of private grazing lands. It offers opportunities for better land management, erosion reduction,

water conservation, wildlife habitat, and improving soil structure.
Nutrient Criteria—Estuarine and Coastal Waters 8-19

Cooperative Extension

State land grant universities and Cooperative Extension play an important role in management

implementation. They have the expertise to research, transfer, and implement agriculture management

systems that will be needed to meet nutrient criteria. In addition, they have developed models and other

predictive management tools that will aid in selecting the most appropriate management activities. 

Contact your local Cooperative Extension Agent or the Agriculture Department at a State land grant

university for more information on the services they can provide.

8.3 COMPREHENSIVE PROCEDURE FOR NUTRIENT MANAGEMENT

Numeric water quality criteria adopted by States into their water quality standards can also serve as

effective scientific tools for comprehensive water and land resource management. Effective programs

incorporate aspects of prevention and maintenance as well as restoration. It is important that existing

high-quality waters be managed wisely as a public resource, and waters whose uses are not yet threatened

or impaired, but nonetheless are at risk from ongoing pollution, should be identified and managed so that

designated uses are maintained in the future. The following 10-step management program that originated

with the Wisconsin Inland Lakes Program (Gibson et al. 1983) has since been refined to become a natural

resources management approach with utility for any water resource. States and Tribes can use this

approach in addition to established regulatory protocols. Though intended to protect or enhance coastal

marine or estuarine waters, the program does not, however, establish or replace any mandated procedures

as part of a regulatory requirement.

Management of tidal and marine waters may be approached as a rational progression of actions beginning

with a statement of major stressors and symptoms and progressing logically to a course of action and

final assessment to determine the relative success of the effort. The following steps illustrate this

management approach. States or communities are encouraged to adapt this technique to suit their

particular needs and expectations. Where considerable information is already available, some of these

steps may be skipped, but the methodology is presented here in detail for consideration.

Step 1: Status Identification

Data used during the preliminary nutrient criteria development process and the application of the criteria

will present the resource manager with the general status of the estuary or coastal area and the need for

responsive action. The information associated with these efforts, however, usually indicates a broad

status condition, for example, high nutrient concentrations, algal blooms, fish kills, loss of seagrasses, or

low-dissolved oxygen. Available data should be evaluated carefully to tease out potential connections to

land use practices or recent changes in practices (e.g., development, fishing pressure, stocking or lack

thereof). In particular, previous investigations should be reviewed to make a preliminary determination

of anthropogenic cause(s) as opposed to natural cycling of the system. Trend assessments on the

Chesapeake Bay, for example, are based on 10-year intervals in an effort to distinguish seasonality from

cultural impacts. Essentially, a preliminary evaluation of readily available data is necessary to ascertain

that there is indeed a problem or potential problem brought on by nutrient overenrichment and that the

sources of the threat probably can be addressed to the betterment of these waters and the public good.
8-20 Nutrient Criteria—Estuarine and Coastal Waters

Step 2: Background Investigation

Given that the initial information reveals a viable management concern, it then becomes necessary and

justifiable to gather as much background information as possible about the waters in question. There are

three primary sources of such information.

Literature Searches

The initial effort here should be a search of the "gray literature" (often internal regional State and Federal

agency reports that provide specific information about the relevant waters). Sources of such information

include natural resource and fisheries agencies, forestry services, water quality administrations,

hydrological and geological survey offices, planning offices, multi-State or county commissions, and

community or environmental groups. A second source to comb through is peer-reviewed professional

literature journals and related publications such as proceedings of conferences and symposiums, which

may include specific studies of the estuary or coastal region of concern. The primary value of this

source, however, probably will be discussions of methods and techniques of investigation and

management. As the management investigation progresses, these sources of information become more

pertinent.

Questionnaires

In preparing a list of agencies from which reports may be solicited, the names of key personal contacts

should emerge. These contacts are the biologists, chemists, specialists, academics, resource managers,

and citizen activists most familiar with these waters. As the literature and baseline data are reviewed,

particular questions should develop, the answers to which will provide a fuller understanding of the

resource and lend direction to the investigation and eventual management plan. Particularly helpful will

be an understanding of the historical antecedents of the present status of the waters.

A standardized questionnaire can be prepared listing concerns such as the availability of any reports or

data or understanding of the history of development in the watershed, perhaps including industries,

agricultural practices, or development and structures associated with the resource. Particular episodes

may be noted for comment in the questionnaire, such as hurricanes, fish kills, algal blooms, or spill

events, as well as historical problems, such as marina development, port facilities, spoils disposal sites,

agricultural runoff, erosion problems, or development concentrations.

All discharge sites should be documented, such as industrial discharges, "Superfund sites," wastewater

treatment plants, drains, concentrations of housing with onsite wastewater treatment, marinas, major road

crossings, and tributaries potentially bearing large loadings of sediments or nutrients. Problem land use

areas along the shore also should be noted, for example, degraded wetlands, embayments where blooms

or fish kills regularly occur, or areas where seagrasses recently have contracted. In addition, it is helpful

to include a large, fairly detailed line drawing of the estuary or portion of the estuary or coastal reach and

its proximal watershed that the respondent may use to reference particular observations.

If at all possible, the questionnaire should be limited to no more than two pages of questions, including

space for answers plus the line drawing. Questions should be direct and concise. Determine exactly
Nutrient Criteria—Estuarine and Coastal Waters 8-21

what you wish to learn and ask questions specifically related to this information. Opportunities for

additional comments should include an open-ended question at the end of the questionnaire.

To get the best response to a questionnaire, the potential respondents should be called first to confirm

their mailing addresses and availability. They should be advised of the nature of the study and their

cooperation then requested. Other potential respondents may be identified through these calls. If a large

survey is necessary, this preliminary step may not be possible. Most regional inquiries, however, are

made usually to no more than 50 specialists, and the additional information gained is well worth the

telephone calls.

Interviews

By this point in the background investigation process, the key people to contact for detailed information

should be evident. Their names will have come up in conversations and on reports, and they will be the

people providing the most helpful responses on the questionnaires.

Other valuable contacts are Basin Commissions; Interstate and State management agencies; EPA,

NOAA, and U.S. Fish and Wildlife Service specialists; the USDA Cooperative Extension Service agents

for the area; State, county, and municipal planners; and university faculty. Long-term residents and

commercial as well as sport fishermen and their organizations should be contacted also. Anecdotal

information can be invaluable and helps add perspective to other sources of data.

The interviews should assess the questionnaire data gathered already; they should clarify and elaborate

on the basic information generated. The interviews also are the means by which apparent contradictions

in perceptions or observations may be resolved at least partially. It should be noted that many people are

uncomfortable with recorded interviews; note-taking often is less intimidating. In either case,

immediately after each interview, a record of answers and observations should be prepared while the

impressions of the interviewer are still fresh.

The information compiled from each background investigation should clarify further the initial problem

stated. It should help resolve any ambiguities about the dynamics of the system and the human

community. In addition, the compilation should identify areas where more definitive, primary data

collection is required to clearly understand the nutrient problems of a particular waterbody and provide

direction for the subsequent management project.

Step 3: Data Gathering and Diagnostic Monitoring

Data obtained during the nutrient criteria development process are the mainstay of the database to be

prepared for any subsequent investigation. The intent of such a process is to develop a reasonable image

of the status of the estuary, bay, or coastal region. Diagnostic monitoring should expand on that structure

and extend the understanding from status of the resource to a diagnosis of causes of the overenrichment. 

For example, where several tributaries of an estuary or salinity zone have been sampled and two are

identified as being of concern, they and other higher order feeder streams must now be sampled to further

target the locations of probable loadings. Though earlier sampling was done to portray the enrichment
8-22 Nutrient Criteria—Estuarine and Coastal Waters

state of the waters, subsequent sampling should focus on near-shore areas of potential loadings,

tributaries, and portions of the tributaries where loadings may originate.

Diagnostic monitoring supports the identification of water quality problems and helps to develop an

appropriate management plan. General guidelines for conducting diagnostic monitoring are as follows. 

Parameters To Sample

Diagnostic monitoring is conducted after nutrient criteria have been established. It might not be

necessary, therefore, to sample some parameters that are not related to the criteria.

• Diagnostic sampling for nutrients requires an estimation of nutrient loading and sources. Major

potential sources of nutrients (e.g., tributary streams, groundwater flow, runoff, illegal discharges,

atmospheric deposition) should be identified and sampled in such a way as to obtain an estimate of

annual loads from each source. 

• The variables and techniques employed in the preliminary survey should be reviewed for adequacy

and either repeated or augmented. A manager should not eliminate the basis of the original

classification by dropping any variables or stations at this point. Documenting potential success or

failure of the subsequent management program will require "before" and "after" databases, and the

initial survey design should be modified only after careful consideration and due attention to

reestablishing the baseline survey.

Flow measurements also are an essential part of this survey. Perhaps more important than in any other

water resource investigation, attention to tidal state and amplitude and seasonal hydrologic characteristics

is imperative to determining the extent and source of nutrient loadings to estuaries and coastal marine

waters. If nutrient concentrations are to be compared meaningfully and loading estimates made,

cross-sectional areas and flow rates for all tributary streams and discharges also must be included in the

survey design. These measurements must be made or extrapolated whenever water quality samples are

collected. Without this information, assigning priorities to various loading sources identified in the

investigation will be difficult or impossible.

Sampling Frequency

Sampling frequency will increase for diagnostic monitoring because the sample population is now a

particular area of the waterbody. Sampling should occur repeatedly during the growing season to

precisely characterize individual areas as well as discharges and loadings. Statistical power analysis can

be used to determine the appropriate sample size based on the purpose of the sampling and the acceptable

error (see Chapter 5).

In addition to expanding the number of stations and parameters to accommodate diagnostic

determinations, the survey design should address temporal variables by sampling these stations during

each season of the year at times calibrated to the particular climate and locale. Accommodation may be

needed for periods of base flow, maximum runoff, turnovers, periods of maximum and minimum
Nutrient Criteria—Estuarine and Coastal Waters 8-23

productivity, and, in some instances, migratory patterns of fish or waterfowl. Seasonal changes in land

use such as peak summer or winter vacation periods, agricultural applications and harvests in the

watershed, and seasonal commercial or industrial activities also should be addressed.

To separate signals from seasonal "noise," it may be necessary to gather survey data for 2 or more

consecutive years to strengthen data assessment (as noted above, some estuary management programs

survey over as many as 10 years, requiring an extensive operating budget). Such assessments will require

a robust statistical evaluation of the data; this element should be incorporated into the study design at the

outset. As with the initial survey design, the preliminary statistical tools chosen may be carried into this

subsequent design as well. Care should be taken to replicate sample collections to ensure representative

sample design and confidence in the results obtained. Early inclusion of a skilled environmental

statistician on the management team is advisable.

Sampling Location

If turbidity, nutrients, and algae are known to be variable across the surface of an estuary or salinity zone,

then multiple sample sites within that zone are required. The exact number of sampling sites in a zone is

determined by the spatial variability of nutrients, turbidity, and chlorophyll and the desired precision. In

general, within a basin or zone, variation in time is larger than variation in space (Knowlton and Jones

1989). Thus, chlorophyll samples taken 2 weeks apart may differ severalfold, but samples taken on the

same day 500 meters apart are likely to differ much less. Depending on the questions being addressed in

the investigation, spatially composite samples may be more cost-effective than separate samples from

several sites.

The design and placement of these sample stations will rely heavily on the proximal and watershed land

use information garnered from the background investigation. The overall objective should be to bracket

suspected sources of nutrient loadings in the tributaries and near-bank areas so parcels can be either

selected or eliminated as potential candidates for management attention.

Step 4: Source Identification

The cumulative information gathered should now provide a clear image of the state of the estuarine or

coastal segment, the most likely sources of nutrient loadings or related degradation, and their relative

contributions to the problem. It is important to note that this process reveals only local sources of the

overenrichment. Atmospheric deposition of nitrogen compounds and other broad-scale impacts beyond

the watershed scale are not specifically addressed and must be assumed as essentially an environmental

constant. With all the risks this constant entails, it is probably not an undue assumption; remediation of

such depositions is probably beyond the scope of most nutrient management projects employing this

guidance. 

The problems to be identified are likely to be as diverse as the geology, hydrology, and land use practices

of the waterbody and watershed. Typical developments include sediment resuspension and nutrient

re-release; biotic imbalances affecting nutrient utilization caused by overfishing or stock

mismanagement; discharge of excess nutrients directly to the waters by wastewater treatment plants,
8-24 Nutrient Criteria—Estuarine and Coastal Waters

storm water runoff, or failing septic systems; and runoff from municipalities, subdivisions, farms,

commercial enterprises, and industrial activities. Other problems have included concentrations of

migratory and resident waterfowl contributing to an excess of nutrients, removal or filling of bank areas

and wetlands that once intercepted nutrient runoff, runoff of herbicide applications that killed

macrophytes and promoted nuisance algal blooms, and chronic, low-dissolved oxygen problems attendant

to overenrichment and vegetative imbalances. 

Any combination of these in situ and land use problems can cause a cumulative overenrichment problem. 

Management planning requires identifying first the loading sources and, second, of those sources, the

ones that are most significant. Proximity of a source to a lake or reservoir (or in some cases, the

ubiquitous nature of a source throughout a watershed such as subdivision or farm runoff), the relative

loading estimate of that source, and the likelihood of successful remediation are the key factors in

deciding which problem sources are priorities for inclusion in a management plan.

Loading estimation models are valuable for estimating the relative significance of various nutrient

sources in the watershed with respect to the likely response of the waters. Chapter 9 describes many of

these models and their relative utility. Modeling permits a manager to try out various scenarios and

combinations of techniques to estimate their likely effectiveness. Some of these options for a nutrient

management plan are discussed in the next step. 

Step 5: Management Practices for Nutrient Control

Once the major sources of concern are identified and agreed on by the management planners, remedial

measures appropriate to these sources must be identified. Management practices are well defined and

documented for a variety of land uses in EPA guidance documents, USDA manuals, U.S. Forest Service

manuals, and urban land use planning guides. Resource managers should study these references for

likely approaches to consider and then consult regional experts in each of the subject land uses for

qualification and other suggested management practice recommendations. Bringing these specialists

together as a small workgroup is an effective, although sometimes contentious, way to develop the most

technically sound approaches to such problems.

Fitting the various components together in a comprehensive management plan is challenging. It calls for

both imagination and cooperation. Usually no one approach stands out as the obvious best choice. 

Instead, two or three permutations of several generally agreed on BMPs will evolve from the planning

sessions.

Selection of the optimal approach—or more likely, the best candidates—should first involve careful

assessment by the planning workgroup and then consultation with all elements of the watershed

community, both organized interest groups and private landowners. The first phase should be conducted

using the threefold framework of evaluation developed by the Department of Resource Development at

Michigan State University (Figure 8-2). The premise behind this approach holds that the most effective

and achievable management plan should address three elements of practicality: scientific validity,

sociopolitical consideration, and economic consideration.
Nutrient Criteria—Estuarine and Coastal Waters 8-25

Figure 8-2. “Threefold framework” of evaluation.

• No resource or environmental management plan should be considered unless it is scientifically valid. 

The technology proposed should be based in sound science and tested and validated. No attempt to

manipulate the environment and peoples' land use prerogatives should be made unless it can be

demonstrated in advance that the technique is reliable or at least that the risks are quantifiable and

understandable.

• The proposed approach should be cost-effective and affordable by the community. Among

technically sound plans to achieve desired goals, the most cost-effective (typically those with

elements that have the greatest benefit-cost ratios) are the easiest to implement and most likely to

satisfy the public interest.

• The management plan should have adequate social and political acceptability. A plan that seems

rational and cost-effective may conflict with the collective values of the local public. Any action

taken in addition to existing requirements should always be researched carefully for justification,

efficacy, lead time required, and likely effects on various segments of the community. 

The resource manager most likely to achieve success will consider and responsibly address each of these

three elements. All candidate alternatives should be evaluated in this manner and revised as necessary. 

Such evaluation not only generates the optimal plan (or plans where competing but different strengths are

evident), but documents the rationale, essential for public review before the final selection is made.
8-26 Nutrient Criteria—Estuarine and Coastal Waters

Involving the public in the process throughout is highly beneficial, and invitations to meetings or

advisories to all potentially interested parties should be provided regularly, if not from the outset then

certainly before plan selection and approval are needed. A balance must be struck between making

public announcements too early, which could arouse people before sufficient information has been

generated, and making announcements too late, which may lead to suspicions that information is being

withheld from the public.

Step 6: Detailed Management Plan Development

A detailed management plan should include all 10 steps of the process described here. The first five

steps are necessary to achieve the design of the plan, but they also should be included so that anyone

reading it will understand what has gone into the effort.

Natural resource management efforts can include one or more of these three elements: education,

financing, and regulation; it is best to initiate them in the order presented here. Start with relatively

low-cost information and education efforts to acquaint people with the problem and how you propose to

address it, and to obtain their suggestions and perceptions. A good educational effort should be the

incentive for volunteer agreements and cooperative action. A grant-in-aid or other assistance often is the

key element to encourage individuals or jurisdictions to adopt appropriate water resource protection

practices. Regulatory actions are necessary and appropriate when mandated by law, when cooperation

and compliance are unlikely to occur otherwise, and when voluntary efforts have not succeeded.

Step 7: Implementation and Communication

Periodic progress reports during implementation of the management project are opportunities to

communicate with administrators, other involved agencies, politicians interested in the project, the

general public and landowners, and other interest groups. Progress reports should be brief and candid. 

They will be part of the public record so that all parties are properly informed, the post-project cry of

inadequate notification is avoided or at least minimized, and techniques and methods used are

documented.

Regional public meetings and hearings are excellent ways to communicate. The more controversial an

issue, the more communication is necessary. To continue implementing a plan despite significant

opposition without evaluating the consequences is a mistake, especially if a change to a step in the

management plan with additional public consultation would still achieve the same objective.

Step 8: Evaluation Monitoring and Periodic Review

The management plan should always include "before, during, and after" water resource quality

monitoring to demonstrate the responses of the system to management efforts. This is the reason for

maintaining and expanding the initial survey stations. Monitoring data are important for evaluating

progress and are included in the requisite progress reports described above. The change or lack thereof

in the status of the estuarine or coastal waters is the ultimate determination of management success.
Nutrient Criteria—Estuarine and Coastal Waters 8-27

These built-in monitoring schedules should include seasonality and periodic data assessment intervals for

management review to permit responses to changing circumstances, modifications of methods, schedules,

and changes of emphasis as needed.

Step 9: Completion and Evaluation

Management projects are frequently planned, initiated, and concluded and even new initiatives

undertaken to meet pressing schedules without sufficient evaluation of the initial project.

Reviews of progress reports, of the original objectives, and of the monitoring data will reveal whether the

lake or reservoir trophic state was successfully protected or improved. Evaluation also provides

documentation for determining whether the project's methods and techniques can be applied elsewhere,

perhaps with modification. Finally, it will reveal any mistakes that should be avoided in future projects

and perhaps will demonstrate that a sequel project is required to fully accomplish the original objectives.

Step 10: Continued Monitoring of the System

The monitoring initiated and expanded in the course of the project can now be reduced to periodic

measuring of key variables at critical times and locations. The purpose now is to keep sufficiently

informed of the status of the waters to ensure that the protection or remediation achieved is maintained. 

If periodic evaluation monitoring indicates a return of trophic decline, intervention should be possible at

an early point so that costs of preserving that which was achieved are reduced. These evaluation and

periodic monitoring steps essentially complete the process. If new issues arise, the manager returns to

Step 1 with a new problem statement. General guidelines associated with evaluation monitoring are

provided below.

Parameters To Sample

Each water quality parameter discussed in the Indicators chapter—TP, TN, chlorophyll a, Secchi depth,

and dissolved oxygen as well as perhaps selenium and vegetation and indicator organisms—should be

sampled during maintenance monitoring. Because the purpose of maintenance monitoring is to

determine if conditions have changed or if criteria are exceeded, other physical or chemical variables

need not be measured.

Sampling Frequency

Sampling efforts for maintenance monitoring can be adaptive and sequential, so that a certain minimum

of information is collected at regular intervals, and if data indicate change or uncertainty, the sampling

effort (in both time and space) can be increased to attempt to reduce the uncertainty. For example, a

station in an undisturbed area could be sampled once every 5 years, from a single visit during an index

period (say, midsummer). If results suggest a change in conditions beyond what is normally expected for

these waters, then additional and more frequent sampling can be continued to determine if the departure

from "normal" conditions is real and if it is ecologically significant. If TP, TN, chlorophyll a, and Secchi

depth relationships have been established, it may be cost-effective to use Secchi depth as a preliminary

indicator; if a trigger value is detected, more parameters can be measured.
8-28 Nutrient Criteria—Estuarine and Coastal Waters

Such variation also suggests different levels of maintenance monitoring, depending on existing

knowledge of the waters and expectations. Maintenance monitoring may be done for several purposes:

• Routine monitoring of waters of known quality (i.e., sampled before) that are not expected to change

greatly

• Initial sampling of a station of unknown quality

• Monitoring of a station or stations of known quality expected to change, say, as the result of

watershed development or restoration efforts

Routine monitoring of stations of known quality is the least intensive and typically requires sampling

once every several years, as in the example above. However, initial sampling of an estuary or coastal

area of unknown quality requires the same sampling effort, and parameters, as the classification survey. 

Monitoring a known estuary or portion of an estuary or coastal area that is expected to change or

suspected to have changed requires more intensive effort, typically an increase in sampling frequency to

several times during the growing season, to obtain seasonal averages of indicator values.

The actual frequency of sampling should be determined by the number of samples required to detect an

ecologically relevant change in the indicators, resources available for the monitoring program, and

amount of time for a change to be detected. These considerations require power analysis using existing

or preliminary data, and tradeoffs of desired significance level, desired power, desired effect size that is

detectable, ecological significance, and most important, resources (labor and money) available for the

monitoring program.

Sampling Location

For routine monitoring, it is recommended that the sampling locations be at least the same as for the

classification survey to provide the database with a certain minimum continuity.

8.4 RESOURCES

Listed below are selected publications concerning coastal or estuarine and watershed management and

protection.

• National Research Council. 2000. Clean Coastal Waters - Understanding and Reducing the Effects

of Nutrient Pollution. National Academy Press. Washington, DC.

• Gibson, G. R., M.L. Bowman, J. Gerritsen, and B.D. Snyder. 2000. Estuarine and Coastal Marine

Water Bioassessment and Biocriteria Technical Guidance. EPA 822-B-00-024. U. S. Environmental

Protection Agency; Office of Water; Washington, DC. 
Nutrient Criteria—Estuarine and Coastal Waters 8-29

• Sharpley, AN, ed. 2000. Agriculture and phosphorus management: the Chesapeake Bay. Boca

Raton, FL: Lewis Publishers.

This text is a compilation of conference proceedings describing nutrient dynamics in the watershed

of Chesapeake Bay, with emphasis on agricultural loadings and practices. Although directed at an

estuarine environment, much of the agriculturally based nutrient information has broad application.

• U.S. Environmental Protection Agency. 1993. Guidance Specifying Management Measures for

Sources of Nonpoint Pollution in Coastal Waters. EPA-840-B-92-002.

The EPA Office of Water produced the 1993 guidance document to support the Coastal Zone Act

Reauthorization Amendments of 1990. This document describes several management measures to

control nonpoint sources of pollution, including nutrients. 

• U.S. Environmental Protection Agency. 1995. Watershed Protection: A Project Focus. EPA

841-R-95-003.

This document focuses on developing watershed-specific programs or projects. It provides a

blueprint for designing and implementing watershed projects, including references and case studies

for specific elements of the process. The document illustrates how the broader principles of

watershed management, including all relevant Federal, State, Tribal, local, and private activities, can

be brought to bear on water quality and ecological concerns.

• U.S. Environmental Protection Agency. 1995. Watershed Protection: A Statewide Approach. EPA

841-R-95-004. 

This document is primarily designed for State water quality managers. A common framework for a

statewide watershed approach focuses on organizing and managing a State's major watersheds (called

basins in this document). In this statewide approach, activities such as water quality monitoring,

planning, and permitting are coordinated for multiple agencies on a set schedule within large

watersheds or basins.

• U.S. Environmental Protection Agency. 1997. Monitoring Consortiums: A Cost-Effective Means to

Enhancing Watershed Data Collection and Analysis. EPA 841-R-97-006.

This document addresses coordination in watershed monitoring. As demonstrated in the document's

four case studies, consortiums can stretch the monitoring dollar, improve cooperation among

partners, and increase sharing of expertise as well as expenses of data collection and management.

• U.S. Environmental Protection Agency. 1997. Land Cover Digital Data Directory for the United

States. EPA 841-B-97-005.
8-30 Nutrient Criteria—Estuarine and Coastal Waters

Land cover, which is the pattern of ecological resources and human activities dominating different

areas of the Earth's surface, is one of the most important data sources used in watershed analysis and

the management of water resources throughout the country. The 75 land cover data summaries in

this directory include contact information to assist readers who may want to acquire copies of the

digital data for their own use.

• U.S. Environmental Protection Agency. 1997. Designing an Information Management System for

Watersheds. EPA 841-R-97-005.

This document is an introduction to the information management responsibilities and challenges

facing any watershed group. The document reviews the fundamentals of identifying information

management needs, integrating different databases, evaluating hardware and software options, and

developing implementation plans.

• U.S. Environmental Protection Agency. 1997. Information Management for the Watershed

Approach in the Pacific Northwest. EPA 841-R-97-004.

This document centers on a series of interviews with leaders and key participants in the statewide

watershed approach activities in the State of Washington. The document reviews Washington’s

statewide watershed activities in case study fashion.

• U.S. Environmental Protection Agency. 1998. Inventory of Watershed Training Courses. EPA

841-D-98-001.

This inventory provides one-page summaries of 180 watershed-related training courses offered by

Federal and State agencies; it also lists resource professionals in the private sector.

• U.S. Environmental Protection Agency. 1997. Statewide Watershed Management Facilitation. EPA

841-R-97-011.

This document addresses statewide watershed management and the process of facilitating the

development or reorientation of statewide watershed programs. It includes State case histories.

• U.S. Environmental Protection Agency. 1996. Watershed Approach Framework. EPA

840-S-96-001.

This publication revisits and updates EPA's vision for a watershed approach, first explained in a 1991

document entitled "Watershed Protection Approach Framework." It describes watershed approaches

as coordinating frameworks for environmental management that focus public and private efforts to

address the highest priority problems in defined geographic areas, involving both ground and surface

water flow.
Nutrient Criteria—Estuarine and Coastal Waters 8-31

• U.S. Environmental Protection Agency. 1997. Top 10 Watershed Lessons Learned. EPA

840-F-97-001.

Watershed work has been going on for many years now, and this 60-page document summarizes the

top lessons that have been learned by watershed practitioners across the United States regarding what

works and what does not.

• U.S. Environmental Protection Agency. 1999. Catalog of Federal Funding Sources for Watershed

Protection (second ed.). EPA 841-B-99-003.

Many sources of Federal funding are available to support different aspects of watershed protection

and specific types of local-level watershed projects. This document presents information on 52

Federal funding sources (grants and loans) that may be used to fund a variety of watershed projects.

• U.S. Environmental Protection Agency. 1997. Watershed Training Opportunities. EPA

841-B-97-008.

This is a 22-page booklet developed to highlight watershed training opportunities offered by EPA's

Office of Water and the Watershed Academy. It covers training courses and educational materials on

watersheds produced throughout the EPA Office of Water.

• U.S. Environmental Protection Agency. 1997. Stream Corridor Restoration: Principles, Processes

and Practices. EPA 841-R-97-011. 

This document is a practical reference manual and logical framework to help environmental

managers recognize stream restoration needs and design and implement restoration projects.

• U.S. Environmental Protection Agency. 1997. Protocol for Developing Nutrient TMDLs. EPA

841-B-99-007. 

This protocol is an organizational framework for the TMDL development process for nutrients. It

leads to an understandable and justifiable TMDL.
Nutrient Criteria—Estuarine and Coastal Waters 9-1

Model Identification and Selection

Model Classification

Use of Models for Nutrient Investigation

Management Applications

CHAPTER 9

Use of Models in Nutrient Criteria

Development

All models are wrong. Some models are useful. 

(George E. P. Box)

9.1 INTRODUCTION

This chapter addresses the role of models in nutrient criteria development. It is closely linked to Chapter

5, which addresses database development, sampling designs, and monitoring. One system is said to

model another when the observable variables in the first system vary in the same fashion as the

observable variables in the second (NRC 2000). Chapter 7 of the NRC report goes on to state that

models come in many forms. They may be empirically derived statistical relationships plotted on a

graph, physical analogues (e.g., mesocosms) of the system of interest, analogues of different systems that

have useful parallel relationships of observable factors (e.g., from physics, the flow of water through

pipes to model the flow of electrons through an electrical circuit), or numerical models run on computers

that are based on first principles or empirical relationships. 

Environmental water quality models have several uses (e.g., reduce ecosystem complexity to a

manageable level, improve the scientific basis for development of theory, provide a framework to make

and test predictions, increase understanding of cause-and-effect relationships, and improve assessment of

factor interaction). Reliable predictions stand out as a salient requirement because of the social and

economic consequences if predictions are unreliable. Many times decisionmakers rely on models to

guide their environmental management choices, especially when costly decisions are involved and the

problem and solution involve complex relationships. This is exemplified by the decision of the Long

Island Sound Hypoxia Management Conference (see Case Study for Long Island Sound). Generally,

empirical and mathematical models are the most widely used models that statistically or mathematically

relate nutrient loads or concentrations to important ecological response variables (e.g., dissolved oxygen

deficiency, algal blooms and related decrease in water clarity, and loss of seagrasses). They both depend

on the scientific robustness and accuracy of underlying conceptual models. 

This chapter addresses both empirical and mathematical models. Considerably more space is devoted to

mathematical models, because they are capable of addressing many more details of underlying processes

when properly calibrated and validated. They also tend to be more useful forecasting (extrapolation)

tools than simpler models, because they tend to include a greater representation of the physics, chemistry,

and biology of the physical system being modeled (NRC 2000). A great danger in complex mathematical

models is that error propagation is difficult to explicitly measure, and there is a tendency to use a more

complex model than required, which drives costs up substantially and unnecessarily. Another

consideration that is gaining acceptance is that mathematical models need to be appropriately scaled to
9-2 Nutrient Criteria—Estuarine and Coastal Waters

spatial and temporal processes, or they may suffer problems similar to empirical models when one

extrapolates the results of scaled experiments to full-sized natural systems. Also, empirical coefficients

introduced into equations often hide the degree of uncertainty concerning the fundamental nature of the

processes being represented. 

Use of Empirical Models in Nutrient Criteria Development

Statistical models are empirical and are derived from observations. To be useful as predictive tools,

relationships must have a basis in our understanding, typically represented by conceptual models. 

However, extrapolation from empirical data is known to be uncertain. Thus, these models are most

reliable when used within the range of observations used to construct the model. When shown to meet

program objectives and requirements, empirical models are a desirable place to begin model development

and, if later determined to be required, they often provide insights into the structure needed for

development of mathematical models. Empirical models typically are useful if only a subsystem of the

larger ecosystem is of primary interest. 

Frequently, the impression is given that the only credible water quality modeling approach is that of

mathematical process-based dynamic computer modeling. This is not the case. For example, a Tampa

Bay water quality modeling workshop in 1992 (Martin et al. 1996) produced the consensus

recommendation that a multipronged (mechanistic and empirical) modeling approach be implemented to

provide technical support for the water quality management process. The Tampa Bay National Estuary

Program produced an empirical regression-based water quality model. The estimated N loads were

related to observed chlorophyll concentrations using the regression model (Janicki and Wade 1996):

C t,s =  t,s +  s * L t, s

where C t,s = average chlorophyll a concentration at month t and segment s,

 L t, s = total N load at month t and segment s,

 t,s and  s are regression parameters.

A related model equated Secchi depth to average chlorophyll a concentrations (Greening et al. 1997). 

This analysis was followed by an empirical model that related N loadings to in-bay chlorophyll a

concentrations.

There are many other examples of empirical models used to relate environmental forcing functions to

ecological responses, especially nutrient load/concentration and response relationships. Much of the

professional aquatic ecological literature reports on use of empirical models (e.g., Chapters 2 and 3). 

Empirical models have their limitations, but when judiciously applied, they offer a highly useful tool to

water quality managers.
Nutrient Criteria—Estuarine and Coastal Waters 9-3

Use of Mathematical Models in Nutrient Criteria Development

Mathematical models can play an important role in assessing acceptable nutrient loads and

concentrations in estuaries and near-coastal areas. For example, models are used to:

• Develop a relationship between external nutrient loads and resulting nutrient concentrations, which

can then be used to define allowable loads

• Define the relationship between nutrient concentrations and other endpoints of concern, such as

biomass or dissolved oxygen

• Provide an increased understanding of the factors affecting nutrient concentrations, such as the

relative importance of point and nonpoint source loads

• Simulate relationships between light attenuation and expected depth of sea grass growth

The intent of this section is to describe the models available for assessing the relationship between

nutrient loading and nutrient-related water quality criteria for estuaries and near-coastal waters. This

chapter provides general guidance and some specific procedures for selecting and applying an

appropriate model. It is divided into the following sections: (1) Model Identification and Selection, (2)

Model Classification, (3) Use of Models for Nutrient Investigation, and (4) Management Applications.

Extensive EPA guidance (i.e., U.S. EPA 1985, 1990a-c; 1997; EPA document # 841-B-97-006) currently

exists on these topics. This section serves primarily to condense the existing guidance with some

modifications, to reflect changes in the science that have occurred subsequent to their publication. In

addition, emphasis is placed on the simpler, more empirical techniques that are applied most easily. 

Readers are referred to the original guidance materials for more detailed discussions of the concepts

described in this section.

9.2 MODEL IDENTIFICATION AND SELECTION

The first steps in the modeling process are model identification and selection. The goals are to identify

the simplest model(s) that addresses all of the important phenomena affecting the water quality problems,

and to select from those the most useful analytical formula or computer model. Selection of too simple a

model can result in predictions of future water quality that are too uncertain to achieve the decisions or

objectives of the study. On the other hand, selection of an overly complex model may also result in

misdirected study resources, delays in the study, and increased cost. Predictive uncertainty may increase

to unacceptable levels because of model parameters that cannot be adequately estimated with available

data. Study costs will increase because of the additional data requirements and the expanded computer

and staff time needed for model runs, analysis, and sensitivity studies. 
9-4 Nutrient Criteria—Estuarine and Coastal Waters

Model Identification

Model identification entails four basic steps:

• Establish study objectives and constraints

• Determine water quality pollutant interactions

• Determine spatial extent and resolution

• Determine temporal extent and resolution

Each is discussed below.

Study Objectives and Constraints

The first step in identifying an appropriate model for a particular site is to clearly delineate the objectives

of the modeling analysis. These objectives address questions such as:

• What are the nutrients of concern?

• What are the environmental endpoints of concern?

• What spatial and temporal scales are adequate for management concern?

• What management issues must the model address?

• What is the acceptable level of uncertainty in model predictions?

The nutrients of concern addressed in this document are nitrogen or phosphorus (depending on which is

the limiting nutrient or will become limiting after controls are implemented). Environmental endpoints

of concern are total nutrient concentration and other indicators of excessive nutrients such as

chlorophyll/biomass and minimum dissolved oxygen. Local, State, and Federal regulations contribute to

the definition of objectives by specifying time and space scales that the model must address: for

example, the averaging period, or the season at which the criteria are applicable. 

All expected uses of the model are to be stated clearly in advance. If the model will be used to predict

future allowable nutrient loads, the specific conditions to be evaluated must be known. Then a final

study objective is established that pertains to the required degree of reliability of model predictions,

which may vary depending on whether the model application is designed for screening level estimation

or for more detailed predictions.

The reliability objective is directly related to project constraints, as there is often a mismatch between

desired model reliability and available resources. Resource constraints can cover four areas: data, time,

level of effort, and expertise. Appropriate model selection must be balanced between competing

demands. Management objectives typically favor a high degree of model reliability, but resource

constraints generally prohibit the degree of reliability desired. Decisions often are required regarding

whether to proceed with a higher-than-desired level of uncertainty, or to postpone modeling until

additional resources can be obtained.
Nutrient Criteria—Estuarine and Coastal Waters 9-5

Water Quality/Pollutant Interactions

After the pollutants and water quality indicators are identified, the significant water quality processes

must be determined. These processes directly or indirectly link the pollutants to be controlled with the

primary water quality indicators. All other interacting water quality constituents thought to be significant

should be included at this point. This consolidation can best be seen in a diagram or flow chart

representing the mass transport and transformations of water quality constituents in a defined segment of

water. Figure 9-1 illustrates variables and processes important to the eutrophication process. Not all of

these need to be included in the actual model selected for use. Those excluded, however, should be

considered externally and reflected in the coefficients. 

At the end of this step all the available knowledge of a system should be assimilated in a way that permits

major water quality processes and ecological relationships to be evaluated for inclusion in the numerical

model description. This conceptual model is the starting point from which systematic reductions in

complexity can be identified to provide an adequate representation of the system while meeting the

objectives of the study.

The simplest level of model complexity considers only total nutrient concentrations and assumes that all

of the processes shown in Figure 9-1 either have no effect on total nutrient concentrations (as is

sometimes assumed for total nitrogen), or can be lumped into a single overall loss coefficient.

Models that simulate phytoplankton concentrations or dissolved oxygen typically include all of the

processes shown in Figure 9-1, and sometimes many more, to describe such processes as sediment

diagenesis and competition among multiple phytoplankton classes. Denitrification in the model is

expressed in terms of the water column carbonaceous biochemical oxygen demand (CBOD).

Spatial Extent and Scale

Two spatial considerations must be addressed in the model identification process: spatial extent and

scale. Spatial extent pertains to the specific boundaries of the area to be assessed. Spatial scale pertains

to the number of dimensions to be considered and the degree of resolution to be provided in each

dimension.

Several guidelines can help locate proper model boundaries. In general, the boundaries should be located

beyond the influence of the discharge(s) being evaluated. Otherwise, proper specification of boundary

concentrations for model projections is very difficult. Boundaries should be located where flow or stage

and water quality are well monitored. Upstream boundaries should be located at a fall line, or at a gaging

station in free-flowing, riverine reaches. Downstream boundaries are best located at the mouth of an

estuary, or even nearby in the ocean. For large estuaries with relatively unaffected seaward reaches, the

downstream boundary can be located within the estuary near a tidal gage and water quality monitoring

station.
9-6 Nutrient Criteria—Estuarine and Coastal Waters

Denitrification 

Settling 

Settling 

Settling 

SOD Settling 

Respiration / Decay 

Water Column 

External Source 

Loads 

Reaeration 

Grazing 

Oxidation 

Nitrif ication 

Respiration 

Photosynthesis 

CBOD 

Organic N 

 NH3 NO2 + 

NO3 SRP 

Organic P 

Zooplankton 

Dissolv ed 

Oxy gen 

Phy toplankton 

Sediment 

Adv ection and 

Dispersion 

Boundary 

Conditions Sediment Flux Temperature Light 

Denitrification 

Figure 9-1. Eutrophication model framework : an example of hypoxia-based conceptual framework

for water quality model. Source: Bierman et al. 1994.

Appropriate model spatial scale requires consideration of two factors: (1) the extent to which spatial

gradients in water quality occur and (2) the extent to which these variations need to be considered from a

management perspective. Real estuaries and near-shore waters all exhibit three-dimensional properties. 

There are gradients in hydrodynamic and water quality constituents over length, width, and depth. The

effective dimensionality of an estuary includes only those gradients that affect the water quality analysis

significantly. 

One-dimensional models consider the change in pollutant concentration over a single dimension in space,

typically oriented longitudinally down the length of an estuary. Two-dimensional models can consider

concentration gradients in the lateral and longitudinal directions (termed x-y orientation), or

concentration gradients that occur longitudinally and vertically (termed x-z orientation). Threedimensional models describe changes in concentration that occur over all three spatial dimensions. 

These models provide the most detailed assessment of pollutant distribution with respect to a discharge;

they also have the most extensive model input requirements and are the most difficult to apply. 

Justifiable reductions in dimensionality result in savings in model development, simulation, and analysis

costs. Usually the vertical and/or lateral dimension is neglected. Eliminating a dimension from the water

quality analysis implies acceptable uniformity of water quality constituents in that spatial dimension. For

example, use of one-dimensional models implies acceptably small deviations in concentration from the
Nutrient Criteria—Estuarine and Coastal Waters 9-7

cross-sectional mean, both vertically and laterally. This judgment requires understanding both the

transport behavior of estuaries and the specific objectives of the study.

Spatial variations are best determined by plotting observed water quality concentrations versus distance

along the dimensions of concern. If such data are not available, other types of methods are available to

estimate the importance of spatial variations. These are described in U.S. EPA (1990a) and discussed

briefly here. The methods can be divided into three categories:

• Relatively simple desktop methods that compare the stratification potential caused by freshwater

inputs to the mixing potential caused by tides and other currents

• Dye studies that observe the degree of mixing

• Geomorphological classification, which categorizes estuaries and the degree of mixing based on

standard morphological categories (e.g., drowned river valleys)

Two situations exist where the observed spatial variations can be ignored. The first is when the primary

location of water quality concern occurs in an area where these gradients are not important. A good

example would be a nutrient modeling study to consider the impacts of a discharge on phytoplankton. If

it takes 2 miles for a bankside discharge to undergo complete lateral mixing, but the location of

maximum algal density is 5 miles downstream, lateral variability in water quality need not be described

by the model. The second situation where a known gradient need not be modeled is where management

objectives are not concerned with the gradient. Examples of this include water quality standards that are

expressed on a spatially averaged basis.

The choice of spatial scale and layout of the model network requires considerable judgment. Knowledge

of the regulatory problem must be combined with knowledge of the loading, transport, and

transformation processes and an understanding of the model chosen to perform the simulations. 

Competing factors often must be balanced, such as precision and cost, or the better fit of one section of

the network versus another. 

Temporal Extent and Scale

The temporal resolution of water quality models falls into one of two broad categories, steady state or

dynamic (i.e., time-variable). Steady-state models predict pollutant concentrations that are expected to

result from a single set of loading and environmental conditions. Dynamic models predict changes in

water quality over time in response to time-variable loads and environmental conditions. 

Steady-state models are much easier to apply and require considerably fewer resources than dynamic

models. This ease of application makes them the preferred modeling framework when loading to the

system can be assumed to be constant and information on changes in concentration over time is not

required. Potential uses of steady-state models include calculation of seasonal average total nutrient

concentrations in response to seasonal average loads. Steady-state models also have been used to predict
9-8 Nutrient Criteria—Estuarine and Coastal Waters

“critical condition” low-flow phytoplankton densities. Steady-state models are poorly suited for

evaluating individual intermittent events (i.e., runoff) but can be used to evaluate the cumulative effect of

multiple, intermittent events on a seasonal basis. 

The timeframe to be represented for a particular steady-state simulation (e.g., monthly average, seasonal

average) must be longer than the water residence time (flushing time) of the estuary. The water residence

time is the time required to remove a parcel of water from an upstream location in an estuary. Factors

that control flushing include tidal action, freshwater inflow, and wind stress. Typical flushing times

range from days in small estuaries, or those dominated by tributary flow, to months in large estuaries

during low tributary flow conditions. Several formulae have been used to estimate flushing times. The

Fraction of Fresh Water Method, the Tidal Prism Method, and the Modified Tidal Prism Method are

fully discussed in Mills et al. (1985) and briefly described in the following section.

Dynamic models should be used when information on changes in concentration over time is required. 

Dynamic models can be divided into two categories, quasi-dynamic and fully dynamic. Quasi-dynamic

simulations predict variations on the order of days to months. The effects of tidal transport are timeaveraged, and net or residual flows are used. Fully dynamic simulations predict hour-to-hour variations

caused by tidal transport. 

The duration of dynamic simulations can range from days to years, depending on the size and transport

characteristics of the study area, the reaction kinetics and forcing functions of the water quality

constituents, and the strategy for relating simulation results to the regulatory requirements. One basic

guideline applies in all cases: the simulations should be long enough to eliminate the effect of initial

conditions on important water quality constituents at critical locations. Flushing times provide the

minimum duration for simulations of dissolved, nonreactive pollutants. The annual sunlight and

temperature cycles almost always require that eutrophication simulations range from seasons to years.

Predicting the year-to-year eutrophication response of large estuaries is best accomplished by quasidynamic simulations. In general, if the regulatory need or kinetic response is on the order of hours, then

fully dynamic simulations are required; if regulatory needs are long-term averages and the kinetic

response is on the order of seasons to years, then quasi-dynamic or steady-state simulations are preferred.

Model Selection

The goal of model selection is to obtain a simulation model that effectively meets all study objectives. In

the final analysis, how a model is used is more important to its success than exactly which model is used. 

Nevertheless, although selection of an appropriate model will not guarantee success, it will help. 

Selection of an inappropriate model will not guarantee failure, but will render a successful outcome more

difficult.

Models may be classified in different and somewhat arbitrary ways. Some models may not quite fit in

any category, or may fit well in several. In addition, models tend to evolve with use. The exact

capabilities of the individual models described here may change. In particular, pollutant fate processes
Nutrient Criteria—Estuarine and Coastal Waters 9-9

may be modified. Usually the computational framework and the basic transport scheme remain stable

over time. For this reason, transport characteristics will provide the basis for the model classification

scheme used here. Models selected for discussion here are general purpose, in the public domain, and

available from or supported by public agencies.

9.3 MODEL CLASSIFICATION

Estuarine and near-coastal models consist of two components: hydrodynamics and water quality. 

Although the hydrodynamic component is independent of the water quality component, water quality

depends on the transport processes controlled by hydrodynamics. As a result, estuarine models can be

classified as Level I to Level IV, according to the temporal and spatial complexity of the hydrodynamic

component of the model. 

Level I includes desktop screening methodologies that calculate seasonal or annual mean total nutrient

concentrations based on steady-state conditions and simplified flushing time estimates. Steady-state

models use an unvarying flow condition that neglects the temporal variability of tidal heights and

currents. These models are designed for relatively simple screening level analyses. They also can be

used to highlight major water quality issues and important data gaps in the early model-identification

stage of a more complex study.

Level II includes computerized steady-state or tidally averaged quasi-dynamic simulation models, which

generally use a box or compartment-type network. Tidally averaged models simulate the net flow over a

tidal cycle. These models cannot predict the variability and range of nutrient concentrations throughout

each tidal cycle, but they are capable of simulating variations in tidally averaged concentrations over

time. Level II models can predict slowly changing seasonal water quality with an effective time

resolution of 2 weeks to 1 month.

Level III includes computerized one-dimensional (1-d) and quasi two-dimensional (2-d) dynamic

simulation models. These real-time models simulate variations in tidal heights and velocities throughout

each tidal cycle. One-dimensional models treat the estuary as well-mixed vertically and laterally. Quasi

2-d models employ a link-node approach that describes water quality in two dimensions (longitudinal and

lateral) through a network of 1-d nodes and channels. Tidal movement is simulated with a separate

hydrodynamic package in these models. The required data and modeling resources typically are

unavailable to support models of Level III or above on a widespread basis.

Level IV consists of computerized 2-d and 3-d dynamic simulation models. Dispersive mixing and

seaward boundary exchanges are treated more realistically than in the Level III 1-d models. These

models are almost never used for routine nutrient assessment, because of excessive resource

requirements.
9-10 Nutrient Criteria—Estuarine and Coastal Waters

Level I Models

Level I desktop methodologies may be employed using a hand-held calculator or computer spreadsheet

and are based on steady-state conditions, first-order decay coefficients, simplified estimates of flushing

time, and seasonal pollutant concentrations. The EPA screening methods provide a series of Level I

analyses as described below. 

EPA Screening Methods

WQAM is a set of steady-state desktop models that includes both one-dimensional and two-dimensional

box model calculations (Mills et al. 1985). Specific techniques contained in WQAM are the Fraction of

Freshwater Method, the Modified Tidal Prism Method, Advection-Dispersion Equations, and Pritchard’s

Two-Dimensional Box Model.

Fraction of Freshwater Method

The Fraction of Freshwater Method estimates pollutant concentrations in one-dimensional estuaries from

information on freshwater and tidal flow by comparing salinity in the estuary with salinity in the local

seawater. The fraction of freshwater at any location in the estuary is calculated by comparing the volume

of freshwater at that location with the total volume of water:

fx (9-1)

S S

S

s x

s

= −

where fx = fraction of freshwater at location x, Ss = seawater salinity at the mouth of the estuary, and Sx =

salinity at location x. 

This ratio can be viewed as the degree of dilution of the freshwater inflow (as well as pollutants) by

seawater. With this in mind, the total dilution of a pollutant input can be calculated by multiplying the

seawater dilution by the freshwater dilution. This then provides a simple way to calculate concentrations

of conservative pollutants. For any location x, at or downstream of the discharge, pollutant loads are

diluted by tidal mixing and upstream flows. The amount of dilution can be calculated by:

C f (9-2) W

Q x x =

where Cx = constituent concentration at location x at or downstream of discharge, fx = fraction of

freshwater at location x, W = waste loading rate (mass/time), and Q = freshwater inflow (volume/time).

The right side of Equation 9-2 can be divided into two distinct terms. The term W/Q represents the

classical equation for determining dilution in rivers caused by upstream flow. The second term, fx,

accounts for the further dilution of the river concentration by tidal influx of seawater.
Nutrient Criteria—Estuarine and Coastal Waters 9-11

Concentrations upstream of the discharge are estimated from the concentration at the point of mix and

the relative salinity of the upstream location. The upstream concentrations are assumed to be diluted by

freshwater to the same degree that salinity is diluted. The equation is:

C f (9-3) W

Q

S

S

x d x

d

=

where fd = fraction of freshwater at discharge location, Sx = salinity at upstream location x, and Sd =

salinity at discharge location.

The fraction of freshwater at the discharge location, fd, is determined by applying Equation 9-1 at the

discharge location. Equation 9-3 can be modified to assess the impact of nutrients entering from the

seaward boundary by replacing the leading fdW/Q term with the boundary nutrient concentration.

Cumulative pollutant impacts from multiple sources are obtained through a two-step process. First,

pollutant concentration caused by each source independent of all other discharges must be determined. 

This determination is accomplished by applying Equation 9-2 or 9-3, one discharge at a time, for any

estuary location of interest. The second step is to determine the total concentration at that location. This

determination is accomplished by adding all of the incremental concentrations caused by each discharge,

as calculated in the first step. This process can be repeated for any location of interest. 

The Fraction of Freshwater Method can be used to predict cumulative impacts in one-dimensional (i.e.,

narrow) systems with significant freshwater inflow. Upstream freshwater flow must be large with

respect to total pollutant inflow for this method to be applicable. The method assumes conservative

pollutant behavior. It is consequently best used to investigate total nitrogen concentrations, because

overall loss rates of total nitrogen from the water column generally are small.

Modified Tidal Prism Method

The Modified Tidal Prism Method estimates dilution from the total amount of water entering an estuary. 

It is more powerful than the Fraction of Freshwater Method because it can consider not only tidal

dilution but also nonconservative reaction losses. It is best applied to investigate total nutrient

concentrations, but provides additional flexibility to describe pollutant losses that may occur through

settling or denitrification. 

The method divides an estuary into segments whose lengths and volumes are calculated using low-tide

volumes and tidal inflow. The tidal prism (i.e., total volume of tidal inflow) is compared for each

segment with the total segment volume to estimate flushing potential in that segment over a tidal cycle. 

The Modified Tidal Prism Method assumes complete mixing of the incoming tidal flow with the water

resident in each segment. 
9-12 Nutrient Criteria—Estuarine and Coastal Waters

The Modified Tidal Prism Method requires seven inputs:

• Freshwater inflow to the estuary

• Salinity of seawater at the downstream boundary

• Pollutant loading rate

• Salinity of each segment

• Low-tide volume for each segment

• Intertidal volume (tidal prism) for each segment

• First-order constituent loss rate for each segment

The first step of the method is to segment the estuary. This requires an initial time-consuming step of

dividing the estuary into segments with lengths equal to the distance traveled by a particle over a tidal

cycle. Cumulative subtidal and intertidal water volumes must be plotted for the estuary, and a graphical

procedure is used to define model segmentation. Once the estuary has been segmented, a series of

calculations can be performed to estimate constituent concentrations in each segment. Specific methods

for dividing the estuary and performing the calculations are provided in Mills et al. (1985).

Advection-Dispersion Equations

Analytical equations have been developed to predict the concentration of nonconservative constituents in

one-dimensional estuaries. These types of equations consider the processes of net seaward flows

(advection) and tidally averaged mixing (dispersion), as well as simple decay. They can be used to

predict total nutrient concentrations at various locations in an estuary in response to alternative nutrient

loading rates. One-dimensional advection-dispersion equations can be expressed in several different

forms (O’Connor 1965), with the most common form contained in the water quality assessment

methodology. These equations require numerous simplifying assumptions, such as constant geometry

and tidal mixing along the length of the estuary, but have proven to be a useful screening tool.

The advection-dispersion equations require five inputs: upstream freshwater flow rate (R), constituent

loading rate (W), estuarine cross-sectional area (A), tidally averaged dispersion coefficient (E), and firstorder decay rate coefficient (k). The first three inputs can be measured directly. The latter two inputs

must be determined indirectly through the model calibration process described below. Two equations are

provided, one which predicts concentrations at any distance (x) upstream of the discharge of concern and

another for concentrations at any distance seaward of the discharge. Co is the concentration at the point

of discharge. The equations are:

C Ce X > 0 (down estuary) (9-4) j x = 0 2

C Ce X < 0 (up estuary) j x = 0 1

where:
Nutrient Criteria—Estuarine and Coastal Waters 9-13

j R

AE

kEA

R 2

2

1 1 4 2

2 = −+

















j R

AE

kEA

R 1

2

1 1 4 2

2 = ++

















C W

R kE U

0

2

1

1 4

=

+ ( /)

These equations can be used to evaluate multiple loading sources by independently applying Equation

9-4 for each loading source and summing the predicted concentrations across the estuary.

Pritchard’s Two-Dimensional Box Model

Vertically stratified estuaries add a significant degree of complexity to the modeling analysis. Pritchard

(1969) developed a relatively simple approach, which can predict nutrient concentration distributions

along the length of an estuary in both an upper and lower layer. This approach is based on numerous

simplifying assumptions, including:

• Steady-state conditions

• Conservative pollutant behavior

• Uniform constituent concentration within each layer or each segment

The following information is required: (1) freshwater flow rate into the head of the estuary, (2) pollutant

mass loading rates, and (3) longitudinal salinity profiles along the length of the estuary in the upper and

lower layers. The method solves a series of linear equations describing the salinity balance around each

segment to determine net flows and dispersion between each segment. Specific methods for performing

the calculations are provided in Mills et al. (1985).

Results from Pritchard’s model can be used to directly calculate conservative constituent concentrations

throughout the estuary or to serve as the hydrodynamic input to one of the Level II models described

below.

Level II Models

Level II models include computerized steady-state and tidally averaged simulation models that generally

use a box or compartment-type network. Steady-state models are difficult to calibrate in situations where

hydrodynamics and pollutant releases vary rapidly. Consequently, these models are less appropriate

when waste load, river inflow, or tidal range vary appreciably with a period close to the flushing time of

the waterbody. Level II models are the simplest models available that are capable of describing the
9-14 Nutrient Criteria—Estuarine and Coastal Waters

relationship between nutrient loads and some of the endpoints of concern of the eutrophication process

(i.e., chlorophyll a, minimum dissolved oxygen).

The Level II models by EPA are QUAL2E and the Water Quality Analysis Simulation Program

(WASP5), with its associated eutrophication program EUTRO5. 

QUAL2E

QUAL2E is a steady-state, one-dimensional model designed for simulating conventional pollutants in

streams and well-mixed lakes (U.S. EPA 1995) and is not recommended for estuaries. Rather, QUAL

TX, which allows tidal boundary conditions, may be more appropriate, but documentation on this model

is very sparse.

WASP6.0

The Water Quality Analysis Simulation Program (WASP6.0) is a general, multidimensional model that

utilizes compartment modeling techniques (Ambrose et al. 1993). The equations solved by WASP6.0 are

based on the principle of conservation of mass. Operated in the quasi-dynamic mode, WASP6.0 requires

the user to supply initial segment volumes, network flow fields, and inflow time functions. The user also

must calibrate dispersion coefficients between compartments. WASP6.0 has the capability of simulating

nutrient-related water quality issues at a wide range of complexity. 

EUTRO5

EUTRO5 is the submodel in the WASP6.0 system that is designed to simulate conventional pollutants. 

EUTRO5 combines a kinetic structure adapted from the Potomac Eutrophication Model with the WASP

transport structure. EUTRO5 predicts DO, carbonaceous BOD, phytoplankton carbon and chlorophyll a,

ammonia, nitrate, organic nitrogen, organic phosphorus, and orthophosphate in the water column and, if

specified, the underlying bed. In addition to segment volumes, flows, and dispersive exchanges, the user

must supply deposition and resuspension velocities for organic solids, inorganic solids, and

phytoplankton. Rate constants and half-saturation coefficients for the various biochemical

transformation reactions must be specified by the user. Finally, the time- and/or space-variable

environmental forcing functions, such as light intensity, light extinction, wind speed, cloud cover,

temperature, and benthic fluxes, must be input.

Level III Models

Level III includes computerized 1-d and 2-d models that simulate variations in tidal height and velocity

throughout each tidal cycle. Level III models enable characterization of phenomena varying rapidly

within each tidal cycle, such as pollutant spills, stormwater runoff, and batch discharges. Level III

models also are deemed appropriate for systems where the tidal boundary impact, as a function of the

hydrodynamics and water quality, is important to the modeled system within a tidal period. 

Tidally varying (intratidal) models have found most use in the analysis of short-term events, in which the

model simulates a period of time anywhere from one tidal cycle to a month. Some seasonal simulations

also have been conducted. 
Nutrient Criteria—Estuarine and Coastal Waters 9-15

In using Level III models, one must decide whether a 1-d longitudinal system is sufficient, or whether a

2-d model is required to capture the longitudinal and lateral variations in the estuary. For estuaries

whose channels are longer than their width and reasonably well mixed across their width, a 1-d model

may be chosen. If large differences exist in water quality from one side of an estuary to the other, or if

vertical stratification is important, then a 2-d model is appropriate.

All Level III models considered here can simulate nutrient-eutrophication interactions. These models

also include settling rates and benthic flux rates for several different constituents, such as phosphorus,

nitrogen, and sediment oxygen demand. The Level III model distributed by EPA is the WASP6. 

Level IV Models

Level IV includes a variety of computerized 2-d and 3-d dynamic simulation models. Dispersive mixing

and seaward boundary exchanges are treated more realistically than in the Level III 1-d models. 

Although not routinely used in nutrient analyses, they are now finding use by experts in special studies. 

Level IV models are required when variations in concentrations in all three dimensions are of concern. 

The time-variable nature of a Level IV model ensures the need for a time-variable watershed model in

order to provide for the nonpoint source inputs. Fully 3-d models that can predict longitudinal, lateral,

and vertical transport are the most complex and expensive to set up and run. 

At present, no Level IV model is supported by EPA. Three current Level IV models, CE-QUAL-W2,

Integrated Compartment Model (ICM), and EFDC, are described below. 

CE-QUAL-W2

CE-QUAL-W2 is a dynamic 2-d (x-z) model developed for stratified waterbodies (Environmental and

Hydraulics Laboratories 1986). This is a U.S. Army Corps of Engineers modification of the Laterally

Averaged Reservoir Model (Edinger and Buchak 1983; Buchak and Edinger 1984a,b). CE-QUAL-W2

consists of directly coupled hydrodynamic and water quality transport models. Hydrodynamic

computations are influenced by variable water density caused by temperature, salinity, and dissolved and

suspended solids. Developed for reservoirs and narrow, stratified estuaries, CE-QUAL-W2 can handle a

branched and/or looped system with flow and/or head boundary conditions. With two dimensions

depicted, point and nonpoint loadings can be distributed spatially. 

CE-QUAL-W2 simulates as many as 20 other quality variables. Primary physical processes included are

surface heat transfer, shortwave and longwave radiation and penetration, convective mixing, wind- and

flow-induced mixing, entrainment of ambient water by pumped-storage inflows, inflow density current

placement, selective withdrawal, and density stratification as influenced by temperature and dissolved

and suspended solids. Major chemical and biological processes in CE-QUAL-W2 include the effects on

DO of atmospheric exchange, photosynthesis, respiration, organic matter decomposition, nitrification,

and chemical oxidation of reduced substances; uptake, excretion, and regeneration of phosphorus and

nitrogen and nitrification-denitrification under aerobic and anaerobic conditions; carbon cycling and

alkalinity-pH-CO2 interactions; trophic relationships for total phytoplankton; accumulation and

decomposition of detritus and organic sediment; and coliform bacteria mortality.
9-16 Nutrient Criteria—Estuarine and Coastal Waters

CH3D-ICM

CH3D is a 3-d, finite-difference hydrodynamic model, developed by the U.S. Army Corps of Engineers

Waterways Experiment Station (WES) in Vicksburg, MS. Results from CH3D have been linked to the

ICM to model water quality in the Chesapeake Bay. The ICM was developed as the integratedcompartment eutrophication model component of the Chesapeake Bay model package. The model

contains detailed eutrophication kinetics, modeling the carbon, nitrogen, phosphorus, silica, and

dissolved oxygen cycles.

CH3D-ICM is a linkage of CH3D, a hydrodynamic model, and ICM, a water quality model. CH3D is a

hydrodynamic model developed for the Chesapeake Bay Program (Johnson et al. 1991). The model can

be used to predict system response to water levels, flow velocities, salinities, temperatures, and the threedimensional velocity field. CH3D makes hydrodynamic computations on a curvilinear or boundary-fitted

platform grid. Deep navigation channels and irregular shorelines can be modeled because of the

boundary-fitted coordinates feature. Vertical turbulence is predicted by the model and is crucial to a

successful simulation of stratification, destratification, and anoxia. A second-order model based on the

assumption of local equilibrium of turbulence is employed. 

ICM is a finite-difference water quality model that may be applied to most waterbodies in one, two, or

three dimensions (Cerco and Cole 1995). The model predicts time-varying concentrations of water

quality constituents and includes advective and dispersive transport. The model also considers sediment

diagenesis benthic exchange. ICM incorporates detailed algorithms for water quality kinetics. 

Interactions among state variables are described in 80 partial-differential equations that employ more

than 140 parameters. An improved finite-difference method is used to solve the mass conservation

equation for each cell in the computational grid and for each state variable.

EFDC

EFDC is a linked three-dimensional, finite-difference hydrodynamic and water quality model developed

at the Virginia Institute of Marine Sciences (Hamrick 1996). EFDC contains extensive water quality

capabilities, including a eutrophication framework based on the ICM model. EFDC is a general-purpose

hydrodynamic and transport model that simulates tidal, density, and wind-driven flow; salinity;

temperature; and sediment transport. Two built-in, full-coupled water quality/eutrophication submodels

are included in the code.

EFDC solves the vertically hydrostatic, free-surface, variable-density, turbulent-averaged equations of

motion and transport; transport equations for turbulence intensity and length scale, salinity, and

temperature in a stretched, vertical coordinate system; and horizontal coordinate systems that may be

Cartesian or curvilinear-orthogonal. Equations describing the transport of suspended sediment, toxic

contaminants, and water quality state variables also are solved.

The model uses a finite-difference scheme with three time levels and an internal-external mode splitting

procedure to achieve separation of the internal shear, or baroclinic, mode from the external free-surface

gravity wave, or barotropic, mode. An implicit external-mode solution is used with simultaneous
Nutrient Criteria—Estuarine and Coastal Waters 9-17

computation of a two-dimensional surface elevation field by a multicolor successive overrelaxation

procedure. The external solution is completed by calculation of the depth-integrated barotropic velocities

using the new surface elevation field. Various options can be used for advective transport, including the

“centered in time and space” scheme and the “forward in time and upwind in space” scheme.

Summary of Model Capabilities

The important features of the models selected for discussion in this manual are summarized in Tables 9-1

and 9-2. The information provided in these tables is primarily qualitative and sufficient to determine

whether a model may be suitable for a particular application. For complete information, consult the

appropriate user's manuals, the supporting agency, and other experienced users.

9.4 USE OF MODELS FOR NUTRIENT INVESTIGATION

This section describes procedures for using models to perform nutrient assessment in estuaries and nearcoastal waters. It describes the model calibration and validation process, where model parameters that

best describe the waterbody of interest are selected. In addition, guidance is provided on using models

for nutrient management and assessment.

 

The first subsection describes a general procedure for calibrating nutrient models, and briefly describes

the validation procedure used to estimate the uncertainty of such models. The subsection also describes

some statistical methods for testing the calibrated models. These methods are useful to aid in the various

calibration phases and also in the validation phase to measure how well model predictions and

measurements of water quality agree.

The second subsection provides guidance on the management application of a calibrated model. 

Methods to project effects of changes in waste loads and to determine causes of existing conditions are

discussed. Finally, a case study application is provided.

Model Calibration and Validation

Model calibration is the process of determining model parameters most appropriate for a given sitespecific application. Calibration of a model involves a comparison of the measured and simulated

receiving water quality conditions. The nature of the model calibration process depends upon the

complexity of the model selected. Simpler models contain relatively few parameters that need to be

calibrated, whereas more complex models contain many.

Calibration alone is not adequate to determine the predictive capability of a model for a particular

estuary. To map out the range of conditions over which the model can be used, one or more additional

independent sets of data are required to determine whether the model is predictively valid. This model

validation exercise defines the limits of usefulness of the calibrated model. Without validation testing,

the model merely describes the conditions defined by the calibration data set. The degree of uncertainty

of any projection or extrapolation of the model remains unknown.
9-18 Nutrient Criteria—Estuarine and Coastal Waters

Table 9-1. Basic model features

Methods/Model

Time

Scales

Spatial

Dimensions

HydrodynamicsData Expertise

Requirements

Distributing

Agency

Scale of

Effort

Fraction of

Freshwater

SS 1D 0 Minimal EPA Days

Modified Tidal

Prism

SS 1D 0 Minimal EPA Days

AdvectionDispersion

Equations

SS 1D 0 Minimal EPA Days

Pritchard’s 2-D

Box Model

SS 2D (xz) 0 Minimal EPA Days

QUAL2E SS 1D I Moderate EPA Few months

WASP5 Q/D 1D, 2D (xy), or

3D

I, S Moderate to

substantial

EPA Few months

CE-QUAL-W2 D 2D (xz) S Substantial Army Corps Several months

CH3D-ICM D 3D S Extreme EPA Months to years

EFDC D 3D S Extreme EPA Months to years

D - dynamic; Q - quasi-dynamic (tidal-averaged); SS - steady state; x - 1-dimensional, xy - 2-dimensional, longitudinal-lateral; 

xz - 2-dimensional, longitudinal-vertical; xyz - 3-dimensional; B - compartment or box 3d; xx - link node branching 2d; 0 - No

hydraulics specified, inferred from salinity data; I - hydrodynamics input; S - hydrodynamics simulated.

In general, models are calibrated in phases, beginning with the selection of the model parameters and

coefficients that are independent of parameters to be calibrated later. For purposes of this discussion, the

process is divided into the categories of hydrodynamic calibration and water quality calibration. The

discussion covers the parameters that need to be calibrated for each level of model as well as the specific

model outputs to be used for the calibration comparison. Calibration of the more complex models

requires detailed guidance; the reader is referred to other documents (e.g., U.S. EPA 1990b; Thomann

and Mueller 1987) for a discussion that is more comprehensive than is feasible here.

Hydrodynamic Calibration

The first phase of calibration concentrates on the hydrodynamic and mass transport models. Two Level I

models, the Fraction of Freshwater Method and the Tidal Prism Method, have no hydrodynamic

parameters that require calibration. In these simplest cases, all hydrodynamic and mass transport

processes are implicitly considered via specification of observed salinity values. Although there are no

parameters to calibrate for these models, there is merit in testing the model’s predictive validity by

comparing predicted concentrations with field observations of a conservative (i.e., nondecaying)

substance, if such data are available.

For the remaining Level I and Level II models, only one hydrodynamic parameter requires calibration: 

the tidal dispersion coefficient. It is possible to calibrate the hydrodynamic and mass transport portions

of these models by determining values for this coefficient that best describe observed salinity or

conservative tracer measurements.
Nutrient Criteria—Estuarine and Coastal Waters 9-19

Table 9-2. Key features of selected models

Model Key Features Advantages Disadvantages/ Limitations

WQAM Simplified equations to

simulate dilution, advection,

dispersion, first-order decay,

empirical relationships

between nutrient loading,

and total nutrient

concentration

Few data requirements;

can be employed easily

with a hand calculator or

computer spreadsheet

Limited to screening and

midlevel applications

QUAL2E Steady-state model provides

adequate simulation of water

quality processes, including

DO-BOD and algal growth

cycles

User-friendly Windows

interface; widely used and

accepted; able to simulate

all of the conventional

pollutants of concern

Limited to simulation of time

periods during which stream

flow and input loads are

essentially constant

WASP5 Based on flexible

compartment modeling

approach; can be applied in

1, 2, or 3 dimensions

Has been widely applied

to estuarine situations;

considers comprehensive

DO and algal processes;

can be used in 3-d

simulations by linking

with hydrodynamic

models

Coupling with multidimensional hydrodynamic

models requires extensive

site-specific linkage efforts

CE-QUAL-W2 Uses an implicit approach to

solve equations of continuity

and momentum; simulates

variations in water quality in

the longitudinal and lateral

directions 

Simulates the onset and

breakdown of vertical

stratification; most

appropriate where vertical

variations are an

important water quality

consideration

Application requires

extensive modeling

experience

CH3D-ICM Finite-difference model can

be applied to most water

bodies in 1 to 3 dimensions; 

predicts time-varying

concentrations of

constituents; includes

advective and dispersive

transport 

State-of-the-science

eutrophication kinetics

Computationally intensive;

requires extensive data for

calibration and verification;

requires a high level of

technical expertise to apply

effectively

EFDC Linked 3-d, finite-difference

hydrodynamic and water

quality model contains

extensive water quality

capabilities; water quality

concentrations can be

predicted in a variety of

formats suitable for analysis

and plotting 

3-d description of water

quality parameters of

concern; entire range of

hydrodynamic, sediment,

eutrophication, and toxic

chemical constituents can

be considered

Computationally intensive;

requires extensive data for

calibration and verification;

requires a high level of

technical expertise to apply

effectively
9-20 Nutrient Criteria—Estuarine and Coastal Waters

Level III models typically contain two calibration parameters, the channel roughness coefficient and the

dispersion coefficient. Occasionally these models are calibrated with current velocity and water surface

elevation data, but more often are indirectly calibrated from salinity or conservative tracer measurements

that also must be used to calibrate the mass transport model. Indirect calibration can result in an

imprecise description of both the circulation and mass transport algorithms, but this is not a severe

drawback unless the critical water quality components of the waste load allocation model are sensitive to

small changes in circulation and mass transport.

Level IV hydrodynamic models contain several calibration parameters, including bottom and surface

friction coefficients; vertical, lateral, and horizontal eddy viscosity coefficients; and wind speed

coefficients. Calibration efforts for these types of models are beyond the scope of this document.

Kinetic Process Calibration

The second phase of calibration involves selection of the set of kinetic coefficients describing nutrient

cycles and other aspects of the eutrophication process. Again, the effort required is directly related to the

complexity of the model selected. 

Two Level I models—the Fraction of Freshwater Method and Pritchard’s model—have no kinetic

parameters that require calibration. The models assume that constituent concentrations undergo no

kinetic processes that affect their concentration, and typically are appropriate only for estimating total

nitrogen concentrations. The remaining Level I models can describe nonconservative constituents, and

lump all kinetic processes into a single overall decay coefficient. Model calibration in these cases

consists of a comparison of predicted versus observed total nutrient concentrations.

The calibration of higher level nutrient and phytoplankton models requires significant expertise because

of the complexity of the interactions between a number of the components of the cycles involved. 

Coefficients that require calibration in these models pertain to: transformation rates among various forms

of a given nutrient; maximum phytoplankton growth rates; phytoplankton respiration rates;

phytoplankton growth sensitivity to light and nutrients; and phytoplankton and detrital settling velocities. 

Model Validation

Validation testing is designed to confirm that the calibrated model is useful at least over the limited range

of conditions defined by the calibration and validation data sets. The procedure is not designed to

validate a model as generally being useful in every estuary, or even as useful over an extensive range of

conditions found in a single estuary. Validation, as employed here, is limited strictly to indicating that

the calibrated model is capable of producing valid results over a limited range of conditions. Those

conditions are defined by the sets of data used to calibrate and validate the model. As a result, it is

important that the calibration and validation data cover the range of conditions over which predictions

are desired. 
Nutrient Criteria—Estuarine and Coastal Waters 9-21

Validation testing is performed with an independent data set collected during a second field study. The

field study may occur before or after the collection of calibration data. For the best results, however, the

validation data should be collected after the model has been calibrated. This schedule of calibration and

validation ensures that the calibration parameters are fully independent of the validation data. Often it is

difficult to assemble the necessary resources to conduct the desired number of surveys. Therefore, it is

important that surveys be scheduled in an innovative manner and the choice of calibration and validation

data sets remain flexible to make the test of the calibrated model as severe as possible. 

Too often, limited studies attempt calibration but not validation. This approach, in effect, limits the

study to describing the conditions during the calibration data collection period and increases the

uncertainty associated with the waste load allocation. In fact, model prediction uncertainty cannot be

reliably assessed in these cases. 

Model Testing

During and after the calibration and validation of a model, at least two types of testing are important. 

First, throughout the calibration procedure, a sensitivity test helps determine which parameters and

coefficients are the most important. Second, a number of statistical tests help define the extent of

agreement between model simulations and measured conditions.

The sensitivity analysis is simply an investigation of how much influence changes in model coefficients

have on simulated results. Typically, important coefficients, parameters, boundary conditions, and initial

conditions are varied by a positive or negative constant percentage to see what effect the change has on

critical predictions. The coefficients and parameters are changed one at a time and the effects typically

are ranked to show which parameters have the most influence and which have the least. 

The second type of testing involves assessment of the “goodness of fit” for model simulations, compared

with measurement of important water quality parameters. In addition to making a visual assessment, a

number of statistical tests have proven useful. These include root mean square error, relative error, and

regression analysis. Other more detailed statistical analyses are described in U.S. EPA (1990b).

The root mean square (rms) error is a criterion that is widely used to evaluate the agreement between

model predictions. The rms error can be defined as:

( ) (9-5) [ ] rms C C N =∑ −m s 2 0.5

where Cm = measured concentration, Cs = simulated concentration, and N = number of measurements.

The rms error can be used to compute simultaneous discrepancies at a number of points, or it can be used

to compute discrepancies between measurements and predictions at a single point over time. Global rms

errors can be computed for a series of measurements at multiple points over time. 
9-22 Nutrient Criteria—Estuarine and Coastal Waters

When discrepancies between model simulations and measurements are not uniform over parts of the

estuary or over time, the relative error may be a more appropriate statistic for testing calibration or

validation. The relative error is defined as:

e (9-6)

C C

C

m s

m

= −

where the overbars denote the average measured or simulated value. Averages can be performed over

multiple sites or over time. The relative error behaves poorly for small values of measurements if

discrepancies are not proportional to the magnitude of the measurement (i.e., small values of Cm magnify

discrepancies) and if Cm > Cs (as the maximum relative error is usually taken to be 100%). Therefore, the

relative error is best for computing composite statistics when discrepancies are not constant, as may

occur when calibration over an extensive range is attempted.

Regression analysis is very useful in identifying various types of bias in predictions of dynamic-state

variables. The regression equation is written as: 

C a bC m s = + +ε (9-7)

where a = intercept value, b = slope of the regression line, and  = the error in measurement mean, Cm.

The standard linear regression statistics computed from the above equation provide insight into the

goodness of fit for a calibration. The square of the correlation coefficient, r2

, measures the percent of the

variance accounted for between measured and predicted values. The slope estimate, b, and intercept, a,

can indicate any consistent biases in the model calibration. A model calibration that perfectly described

all available data would have a correlation coefficient of 1.0, a slope of 1:1, and a zero intercept.

9.5 MANAGEMENT APPLICATIONS

Once the model is calibrated and validated, it can be used to simulate future conditions to determine

effects of changes in waste loads or to investigate causes of existing problems. This section describes

three types of management application: (1) load-response analysis, (2) determination of acceptable

nutrient loads, and (3) investigation of causes of nutrient problems.

Load-Response Analysis

A load-response analysis consists of performing multiple model simulations using different loading rates

and examining the water quality predicted for each simulation. The most common use of a model to

investigate nutrients in estuaries is to determine the water quality throughout an estuary in response to

changes in nutrient loads. Models are designed to predict water quality based on loadings and

environmental conditions (Figure 9-2). 

This type of analysis also requires specification of the environmental conditions (e.g., freshwater inflows,

tidal conditions) to be considered. The results of the load-response analysis are directly related to the
Nutrient Criteria—Estuarine and Coastal Waters 9-23

Figure 9-2. Use of models in load-response analysis.

environmental conditions specified for the model simulation. For example, use of summer-average

environmental conditions in the model will show the response in summer-average water quality to

changes in loads. For the simplest Level I models, environmental conditions are specified implicitly

through the use of salinity observations. Predictions from these models will correspond to the

environmental conditions that were in effect when the salinity was measured. The more complex models

require explicit definition of environmental condition and can be used to provide predictions for a wide

range of environmental conditions.

Acceptable Nutrient Loads

The most common use for water quality models is to define allowable loads necessary to achieve water

quality objectives. As seen in Figure 9-2, models predict water quality for a specified set of loads and

environmental conditions. Determination of acceptable loads typically requires an iterative procedure, as

shown in Figure 9-3. The first iteration consists of performing a model simulation using existing loads

and comparing predicted water quality with objectives. Assuming that the existing loads result in 

unacceptable water quality, additional model simulations are performed using incremental reductions in

nutrient loads until water quality objectives are achieved.

The approach shown in Figure 9-3 can be used to define necessary reductions in total loads as well as

reductions in individual contributors to the total load.

The results of the above approach are highly dependent on the environmental conditions selected, as

allowable loading rates can vary substantially across different environmental conditions. Two

approaches are available for selecting critical conditions for use in defining allowable loads. These

approaches are termed the critical conditions approach and the continuous simulation approach. In the

critical conditions approach, a single set of environmental conditions is selected for analysis. These

conditions typically represent critical or worst-case conditions, that is, those environmental conditions

that will result in the poorest water quality for a given set of loads. The rationale for the critical

conditions approach assumes that if loads are defined to meet water quality objectives during “critical”

conditions, the same loads will result in attaining water quality objectives during most other conditions as

well.
9-24 Nutrient Criteria—Estuarine and Coastal Waters

Figure 9-3. Use of models in determining allowable loads.

The continuous simulation approach performs simulations for as long a duration as is feasible, using

historically observed variations in environmental conditions. The predicted water quality resulting from

the continuous simulation is analyzed to determine the frequency with which water quality problems are

observed to occur.

The overall intent of the modeling analysis is to define loads that will restrict the occurrence of water

quality problems to an acceptable frequency. Each of the above two methods has particular strengths and

weaknesses for performing this task. The continuous simulation approach provides a direct means to

consider frequency of occurrence (i.e., number of years per problem) but has extreme resource

requirements. The ability to perform continuous simulation of sufficient duration typically is constrained

by the availability of data describing historical environmental conditions, or by computational

requirements for the higher level models. The critical conditions approach has much more manageable

resource requirements; however, there are no clear methods to establish appropriate critical conditions. 

In estuaries, freshwater, tides, wind, and other factors all can affect water quality. Selection of

appropriate values for each environmental parameter requires considerable judgment. Furthermore, the

specific level of protection associated with any single set of environmental conditions cannot be

evaluated without performing a continuous simulation.

Case Study Example

Water quality models also can be used to gain an increased understanding of the relative importance of

various loading sources to an estuary or near-coastal water. It is possible to investigate the contribution

of individual loading sources to the water quality problem by performing a series of simulations

examining each loading source separately. Because most water quality models assume a linear

relationship between pollutant load and resulting water quality impact, it is possible to determine overall

impacts to the estuary by summing the impacts from each source.
Nutrient Criteria—Estuarine and Coastal Waters 9-25

13

12

11

10

 9

 8

 7

 6

 5

 4

 3

 2

 1

 City

WWTP

Investigation of Causes of Nutrient Problems

Shipps Creek (Figure 9-4) is a long, narrow, tidal tributary receiving nutrient inputs from upstream runoff

and a single wastewater treatment plant (WWTP). This case study example demonstrates the use of 

models to perform three tasks:

• Estimate the contribution of various loading sources to the overall summer-average total nitrogen

concentration

• Estimate the effect of a 50% reduction in loads from the WWTP on total nitrogen concentrations

throughout the estuary

• Estimate the reduction in loading necessary to achieve an average total nitrogen concentration of

0.100 mg/L in the lower half of the estuary

Short-term answers were required, and screening-level accuracy was judged acceptable because of the

short timeframe and limited data available.

The Fraction of Freshwater Method was selected because the estuary was considered one-dimensional,

long-term average results were acceptable, and the water quality target was specified in terms of total

nutrient levels. 

Figure 9-4. Shipps Creek site map and salinity monitoring location.
9-26 Nutrient Criteria—Estuarine and Coastal Waters

Information available to support the study included summer-average salinity measurements at 13

locations along the length of the estuary, summer-average freshwater inflows and total nitrogen loads to

the estuary, nitrogen loads from the WWTP, and nitrogen and salinity concentrations outside of the

estuary. This information was compiled as shown in Table 9-3 to allow implementation of the Fraction

of Freshwater Method. The top two rows of Table 9-3 show the measured loading rates and seaward

boundary conditions. The two leftmost columns define segments centered around each of the salinity

measurements. The third column applies Equation 9-1 to calculate the fraction of freshwater in each

segment.

The fourth through sixth columns in Table 9-3 apply Equations 9-2 or 9-3 as appropriate to determine the

incremental contribution to the total nitrogen (TN) concentrations from each of three possible

contributing sources: upstream (TNUP), wastewater treatment plant (TNWWTP), and the downstream

seaward boundary (TNSEA). Equation 9-2 is applied only to those segments downstream of the loading

source, whereas Equation 9-3 is applied to those segments upstream of the loading source. Equation 9-2

is applied to all segments for examining impacts from upstream sources, whereas Equation 9-3 is applied

to all segments for examining impacts from the seaward boundary. For determining WWTP impacts,

Equation 9-2 is applied to segments 1-11 and Equation 9-2 is applied to segments 12-13. The final

column in Table 9-3 sums the incremental contributions from each of the sources to provide a prediction

of overall TN concentrations throughout the estuary.

Table 9-3. Calculation spreadsheet for Shipps Creek estuary

Freshwater Inflow Seawater Salinity Upstream Load WWTP Load Seawater TN

Q = 100,000 cmd Ss = 30 ppt W = 5,000 g/day W = 500,000 g/day 0.005 ppm

Segment

#

Salinity, Si

(ppt)

Fraction of

Freshwater, fI

TNUP

(mg/L)

TNWWTP

(mg/L)

TNSEA

(mg/L)

Overall TN

(mg/L)

1 29 0.03 0.002 0.017 0.005 0.023

2 27 0.10 0.005 0.050 0.005 0.060

3 25 0.17 0.008 0.083 0.004 0.096

4 23 0.23 0.012 0.117 0.004 0.132

5 21 0.30 0.015 0.150 0.004 0.169

6 19 0.37 0.018 0.183 0.003 0.205

7 18 0.40 0.020 0.200 0.003 0.223

8 16 0.47 0.023 0.233 0.003 0.259

9 14 0.53 0.027 0.267 0.002 0.296

10 12 0.60 0.030 0.300 0.002 0.332

11 10 0.67 0.033 0.333 0.002 0.368

12 5 0.83 0.042 0.110 0.001 0.152

13 1 0.97 0.048 0.025 0.000 0.074
Nutrient Criteria—Estuarine and Coastal Waters 9-27

Figure 9-5. Model results for existing conditions.

The results of this modeling analysis are shown graphically in Figure 9-5, showing the overall TN

distribution as well as its component. Although the Fraction of Freshwater Method does not require

calibration, it would be worthwhile at this point to confirm that the model predictions of TN throughout

the estuary were consistent with observed data collected over the same time period. Figure 9-5 shows

that the WWTP is the dominant source of nitrogen throughout most of the estuary. Upstream sources are

the dominant component only at the extreme head of the estuary. Nitrogen contributions from the

seaward boundary are small throughout the system. The results in Table 9-2 and Figure 9-4 satisfy the

first objective of this study, which was to determine the contribution of various loading sources to the

overall summer-average total nitrogen concentration. The second objective of the study was to determine

the water quality resulting from a 50% reduction in WWTP TN loads. This was accomplished by

reapplying Equations 9-2 and 9-3 using one-half of the original WWTP loads. Results of this analysis

are shown in Figure 9-6, indicating a decrease in peak TN concentrations from 0.368 to 0.202 mg/L and a

decrease in lower estuary (defined as segments 1-6) average concentrations from 0.111 to 0.064 mg/L. 

This nearly 50% reduction in concentrations was expected, because the original analysis had

demonstrated that the WWTP was the dominant loading source to the estuary.

The final objective of the study was to determine the loading reductions necessary to achieve a lower

estuary average concentration of 0.08 mg/L. No single answer exists to this question, because three

separate sources of nitrogen to the estuary contribute to the total concentration. Analysis of the data in

Table 9-3 shows that the incremental contribution of the upstream, WWTP, and seaward sources to lower

estuary average concentrations were 0.010, 0.100, and 0.004 mg/L, respectively. Because the seaward 
9-28 Nutrient Criteria—Estuarine and Coastal Waters

0.00

0.05

0.10

0.15

0.20

0.25

0.30

0.35

0.40

0.45

0.50

13 11 9 7 5 3 1

Segment Number

Total Nitrogen (mg/l)

Total WWTP

Seaward Upstream

Figure 9-6. Model results for 50% reduction in WWTP load.

boundary nitrogen concentration cannot be controlled, management reductions must be restricted to

either the upstream sources or the WWTP. The WWTP load must be reduced by at least 20% to meet the

target TN concentration of 0.080 mg/L; otherwise, its contribution alone will exceed the target. Beyond

the initial 20% reduction in the WWTP source, further reductions must come either from the WWTP or 

upstream sources. The specific allocation of these load reductions among sources is an economic and

social decision that the model is not designed to address. The model is expressly designed, however, to

test alternative proposals of load reductions to determine if they will meet the water quality objective. 

For example, a 25% reduction in both upstream and WWTP sources resulted in an average concentration

of 0.087 mg/L (i.e., above the target), but a 30% reduction in WWTP loads coupled with a 40%

reduction in upstream loads was shown to just meet the target.

All model simulations presented here should be viewed with extreme caution, because they are based on

an uncalibrated, screening-level model. The level of uncertainty for these predictions cannot be

quantified and is expected to be quite large. The model results do, however, provide the best possible

estimate describing the relationship between nutrient loads and resulting concentrations, given the

available resources.

Overview of Chesapeake Bay Airshed, Watershed, and Estuary Models

The cross-media models used in the Chesapeake Bay analysis consist of three models: an airshed model,

a watershed model, and a model of the Chesapeake estuary. These models are linked so that the output

of one simulation provides input data for another. The simulation period is the 10-year period of January
Nutrient Criteria—Estuarine and Coastal Waters 9-29

1, 1985, to December 31, 1994. Versions of these models have been used by the Chesapeake Bay

Program for more than a decade and have been refined and upgraded several times.

Airshed Model

The Chesapeake Bay Program airshed model provides estimates of atmospheric deposition loads of

nitrogen. A product of the EPA National Exposure Research Laboratory in Research Triangle Park,

North Carolina, RADM (pronounced “radum”) is an acronym for Regional Acid Deposition Model. 

RADM is a three-dimensional model that tracks nutrient emissions across the eastern United States. Two

RADM grids meet various resolution needs. A large grid scale covers the entire RADM domain and

contains 20,000 square cells of 6,400 square kilometers each. A fine grid scale covers the region of the

Chesapeake Bay watershed and has 60,000 cells, each covering 400 square kilometers. The model

domain in the vertical is 15 cells deep, reaching from ground level to the top of the free troposphere. The

depth of the cells increases with altitude. One of the findings of the RADM model is that the Chesapeake

Bay airshed, defined as the area accounting for 75% of the deposition in the watershed, is approximately

5.5 times the size of the watershed.

RADM is used to drive scenarios associated with reductions in atmospheric deposition of nitrogen. A

base condition deposition represents an estimate of the current condition of atmospheric deposition in the

watershed and is developed from a regression of National Atmospheric Deposition Program (NADP)

data. RADM scenarios of atmospheric deposition reductions are incorporated into Watershed Model

scenarios by adjusting the base NADP condition on a segment-by-segment basis with a percent change

prescribed by RADM scenario results. Results from RADM specify loads of wet and dry deposition to

the Chesapeake watershed for the State Implementation Program (SIP) and Limit of Technology

scenarios. Deposition loads are input directly to the land surfaces of the watershed model or to the tidal

water surface of the Chesapeake Bay estuary model package as daily loads of wet deposition (from rain

washout of atmospheric nitrate and ammonia) and 12-year average loads of dry deposition.

Three atmospheric deposition loads were used for the Chesapeake analysis: (1) the base condition of

atmospheric deposition, (2) the estimated atmospheric deposition of nitrogen equivalent to the 1998 SIP

controls of atmospheric deposition, and (3) the estimated atmospheric deposition under full limit of

technology control (see Table 9-4).

Table 9-4. Chesapeake watershed nitrogen deposition under varying management schemes for emissions of

nitrogen atmospheric deposition precursors

Scenario TN Deposition (millions of kg/year)

Base Condition 204

State Implementation Plan (SIP) 178

Limit of Technology 128

Sources: Chesapeake Bay Program Phase IV Watershed Model and RADM.
9-30 Nutrient Criteria—Estuarine and Coastal Waters

Watershed Model

The watershed model simulates nutrient and sediment loads delivered to the Chesapeake Bay from all

areas of the watershed. Land uses of cropland, pasture, urban areas, and forests are simulated on an

hourly time-step, tracing the fate and transport of input nutrient loads from atmospheric deposition,

fertilizers, animal manures, and point sources. The simulation is an overall mass balance of nitrogen and

phosphorus nutrients in the basin, so that the ultimate fate of input nutrients is simulated, either as

incorporation into crop or forest plant material, incorporation into soil, or river runoff. Nitrogen fates

included volatilization into the atmosphere and denitrification. Transport in rivers is simulated to the

tidal waters of the Chesapeake Bay. Sediment is simulated as eroded material washed off land surfaces

and transported to the tidal bay. Scenarios are run for 10 years on a 1-hour time-step, and results are

aggregated into 10-year average loads for comparison among scenarios.

To simulate the delivery of nutrients and sediment to the bay, the watershed was divided into 86 major

model segments, each with an average area of 194,000 hectares. Segmentation, based on three tiers of

criteria, partitioned the watershed into regions of similar characteristics. The first criterion was

segmentation of similar geographic and topographic areas, which were further delineated in terms of soil

type, soil moisture holding capacity, infiltration rates, and uniformity of slope. The second criterion

involved finer segmentation based on spatial patterns of rainfall. Each segment had a bank-full travel

time of about 24–72 hours. The third criterion used to further delineate segments was based on features

of the river reach. River reaches containing a reservoir were separated into a reservoir simulation and a

river simulation of the free-flowing river. For example, the James basin had 11 model segments, 2

represented reservoirs on the James and Appomattox, and the segmentation generally became finer with

closer proximity to tidal waters. 

Model segments were located to take advantage of observed data locations, so that a model segment

outlet was located close to monitoring stations. Water quality and discharge data were collected from

Federal and State agencies, universities, and other organizations. More than 150 subsegments were used

at the interface between the watershed and estuary models to accurately deliver flow, nutrient, and

sediment loads to appropriate areas of the estuary. Increased simulation accuracy motivated the division

of basins into multiple segments and into simulation time-steps of an hour, but all scenario results were

reported at the level of the basin and for 10-year average loads.

The watershed model has been in continuous operation at the Chesapeake Bay Program since 1982 and

has had many upgrades and refinements since that time. The watershed model used for this application

was Phase 4.1 based on the HSPF Version 11 code (Hydrologic Simulation Program - Fortan - HSPF). 

Version 11 is a widely used public domain model supported by EPA, USGS, and the U.S. Army Corps of

Engineers.

Estuary Model

The Chesapeake Bay Estuary Model Package (CBEMP) is actually several models simulating different

aspects of water quality in the bay and tributaries. A water quality model simulates 22 parameters, or

state variables, as listed in Table 9-5.
Nutrient Criteria—Estuarine and Coastal Waters 9-31

Table 9-5. Water quality state variables used in CBEMP

Temperature Dissolved organic nitrogen

Salinity Labile particulate organic nitrogen

Inorganic suspended sediments Refractory particulate organic nitrogen

Diatoms Total phosphorus

Cyanobacteria (blue-green algae) Dissolved organic phosphorus

Other phytoplankton Labile particulate organic phosphorus

Dissolved organic carbon Refractory particulate organic phosphorus

Labile particulate organic carbon Dissolved oxygen

Refractory particulate organic carbon Chemical oxygen demand

Ammonium Dissolved silica

Nitrate + nitrite Particulate biogenic silica

Zooplankton were separated into two size classes: microzooplankton (>44 microns) and

mesozooplankton (>202 microns).

Linked to the water quality model is a hydrodynamic model, simulating the hydrodynamics, or water

movement, throughout the tidal estuary. The hydrodynamic model produced three-dimensional

predictions of velocity, diffusion, surface elevation, salinity, and temperature on an intratidal time scale. 

The model grid of the hydrodynamic and water quality models consists of more than 10,000 cells.

The modeling process involves simulation of living resource parameters, such as dissolved oxygen,

chlorophyll concentrations, and submerged aquatic vegetation (SAV). Computed parameters are

compared with living resource standards, and an estimation is made of the degree to which computed

conditions benefit the resources of interest (e.g., fish, oysters). In addition, the CBEMP includes the

direct interactive simulation of SAV and water quality. Three phytoplankton groups were simulated. 

Over seasonal time scales, the bay sediments are a significant source of dissolved nutrients to the

overlying water column. The role of sediments in the systemwide nutrient budget is especially important

in summer when seasonal low flows diminish riverine nutrient input. In addition, water temperatures

enhance biological processes in the sediments, creating greater sediment oxygen demand. Bay sediments

retain a long-term nutrient load “memory” of several years; that is, sediment nutrient fluxes to the water

column are determined by organic nutrient inputs from several previous years. Therefore, the water

quality model was coupled directly to a predictive benthic-sediment model. These two models interact at

each time-step, with the water quality model delivering settled organic material to the sediment bed, and

the benthic-sediment model calculating the flux of oxygen and nutrients to the water column.

Linked to the CBEMP are the watershed and airshed models, which provide daily input data. Generally,

10-year scenarios are run on 15-minute time-steps with output generated each 10 days. The estuary
9-32 Nutrient Criteria—Estuarine and Coastal Waters

model has been in operation since 1987, with two major model refinements released since the initial 1987

steady-state model.

Further information on the entire suite of Chesapeake Bay Program models, their documentation, and

application can be found at: http://www.chesapeakebay.net/model.htm.
Nutrient Criteria—Estuarine and Coastal Water CS-1

CASE STUDY

SAN FRANCISCO BAY PROGRAM:

MANAGING COASTAL RESOURCES OF THE U.S.

The following case study is extracted from U.S. Geological Survey Fact Sheet FS-053-95 

(available online at: http://water.usgs.gov/wid/html/sfb.html)

Coastal ecosystems, such as bays and estuaries, are among our most disturbed natural environments. 

These ecosystems also are among our most valuable habitats—estuaries supported U.S. fisheries valued

at $19 billion in 1990. Although many human activities cause change in the coastal zone, they occur

against a background of natural change. Effective coastal-zone management requires that we identify

and understand these separate causes of ecosystem change. With this goal in mind, the United States

Geological Survey (USGS) began in 1968 a broad program of scientific study in San Francisco Bay

(Figure 1). The program is based on a conviction that sustained, multifaceted investigation of one

estuary will produce general lessons to guide the management of natural resources associated with all our

coasts. 

The USGS San Francisco Bay Program has produced more than 250 reports, including three books and a

review of the human modifications of the bay. These publications are a source of guidance to resource

managers as they work to understand how human activities (such as water diversion, commercial trade,

and waste inputs) cause change in the coastal zone. The program has been organized around themes. 

One of the most important themes is the integrated study of nutrients, toxic substances, and living

resources at lower levels of the food chain—the phytoplankton and bottom dwelling invertebrates. Close

collaboration between chemists and ecologists has helped to explain how plant and animal species of

coastal ecosystems are organized into food chains, how nutrients and toxic contaminants are incorporated

into these food chains, and how the lessons learned from detailed scientific understanding can be applied

to develop effective monitoring programs and rational environmental standards.

Nutrient Enrichment

Human settlement around coastal water bodies has led to increased inputs of nutrients such as nitrogen

and phosphorus. Many estuaries are now among the most intensively fertilized environments on Earth. 

Each day, San Francisco Bay receives more than 800 million gallons of municipal wastewater containing

60 tons of nitrogen. In response to these concerns, the USGS developed a biological monitoring

procedure that has been used continuously since 1977 near a waste-treatment facility. Monitoring

continued as wastewater-treatment technologies improved. This is the longest continuous record of

contaminant concentrations in a natural environment of the United States. The transfer of monitoring

procedures developed by the USGS to local agencies and businesses serves as a model of cooperation

between research and regulatory agencies.

Management Questions

Water-quality managers need to know how nutrient inputs cause changes in water quality, the natural

capacity of coastal waters to assimilate added nutrients, the level of waste treatment required to protect

living resources from the harmful effects of nutrient enrichment, and if programs of nutrient reduction

are having beneficial effects.

USGS Contributions

Since 1968, the continuous study of San Francisco Bay by the USGS has given that agency a unique

opportunity to follow ecosystem responses to improved wastewater-treatment methods as mandated by

State and Federal legislation. One result of the implementation of these improved methods has been a
CS-2 Nutrient Criteria—Estuarine and Coastal Water

Figure 1. San Francisco Bay has been a focus of intensive investigation by the USGS since 1968.

large reduction in the input of ammonia-nitrogen from some municipal wastewater-treatment facilities

(Figure 2).

USGS studies show that in spite of its nutrient enrichment, San Francisco Bay has not been affected by

harmful algal blooms. This seeming paradox is explained partly by the abundant bottom-dwelling

invertebrates (small clams, mussels, crustaceans) that filter the water and remove new algae as fast as

they are produced. Feeding by these animals is a form of natural waste treatment that helps control the

growth of algae in a nutrient-rich environment.

Concepts and measurement techniques from this USGS program are now incorporated into a locally

funded and managed Regional Monitoring Program.

Lessons Learned

• The chemical quality of coastal waters can respond almost immediately to waste-treatment

improvements.

• Responses of biological communities to these chemical changes can take years or even decades.

• Coastal water bodies have differing sensitivities to waste loading. The most cost-effective national

strategy for regulating nutrient inputs will consider these differences among ecosystems.

For Further Information:

Visit the USGS website on San Francisco Bay at: http://sfbay.wr.usgs.gov/
Nutrient Criteria—Estuarine and Coastal Water CS-3

1970 1975 1980 1985 1990 1995

0

5

1 0

1 5

2 0

2 5

NITROGEN, IN 

MILLIGRAMS PER LITER

YEAR

CONCENTRATION OF AMMONIA

SITE 1

Figure 2. Implementation of advanced wastewater treatment in 1979 immediately reduced the input of

ammonia- nitrogen to South San Francisco Bay. In prior decades, the South Bay had repeated

episodes of oxygen depletion and animal die-offs. USGS measurements have shown a

complete cessation of these episodes since 1980. Spawning salmon have recently been

observed in South Bay streams for the first time since the early 1900's. See figure 1 for

location of site.
CS-4 Nutrient Criteria—Estuarine and Coastal Water

CASE STUDY

LONG ISLAND SOUND - HYPOXIA 

U.S. Environmental Protection Agency

http://www.epa.gov/region01/eco/lis/hypox.html

The Problem

During the summers of 1987-93, from half to two-thirds of the Sound's bottom waters experienced

dissolved oxygen levels below 5 milligrams of oxygen per liter of water (mg/L). Levels of dissolved

oxygen of 5 mg/L and higher are generally accepted as being protective of the Sound's estuarine life. In

1989, a particularly bad summer, more than 500 square miles (40 percent) of the Sound's bottom waters

had dissolved oxygen levels less than 3 mg/L. During most of these years, dissolved oxygen in a portion

of the Sound (up to 50 square miles) fell below 1 mg/L and in 1987 anoxia, the absence of any oxygen,

was recorded in a portion of the Western Narrows. 

These low levels of dissolved oxygen cause significant, adverse ecological effects in the bottom water

habitats of the Sound. To date, research shows that the most severe effects (such as mortality) occur

when dissolved oxygen levels fall below 1.5 mg/L at any time and below 3.5 mg/L in the short-term (i.e.,

4 days), but that there are probably mild effects of hypoxia when dissolved oxygen levels fall below 5

mg/L. The levels regularly observed in the Sound during late summer: 

• Reduce the abundance and diversity of adult finfish;

• Reduce the growth rate of newly-settled lobsters and perhaps juvenile winter flounder;

• Can kill species that cannot move or move slowly, such as lobsters caught in pots and starfish, and

early life stages of species such as bay anchovy, menhaden, cunner, tautog, and sea robin;

• May reduce the resistance to disease of lobsters and other species; and

• Diminish the habitat value of Long Island Sound. 

The Cause of the Problem

Excessive discharges of nitrogen, a nutrient, are the primary cause of hypoxia. Nitrogen fuels the growth

of planktonic algae. The algae die, settle to the bottom of the Sound and decay, using up oxygen in the

process. 

Natural stratification of the Sound's waters occurs during the summer when warmer, fresher water

"floats" on the top of cooler, saltier water that is more dense. This natural stratification forms a density

difference between the two layers called a pycnocline. This prevents mixing of surface and bottom

waters. 

Oxygen from the atmosphere and photosynthesis keep the surface layer well oxygenated, but the oxygen

cannot pass through the pycnocline into the bottom layer very easily. Decaying algae and other organic

material in the sediment and animal respiration in the bottom layer use up oxygen faster than it is

replenished. Hypoxia develops and usually persists as long as the stratification lasts (usually one to two

months in late summer). 

But hypoxia in Long Island Sound is too complex to fully understand using direct observations alone. 

Natural variations in weather and other physical factors affect the extent and severity of hypoxia. The

Management Conference has constructed mathematical models in order to understand the relationship
Nutrient Criteria—Estuarine and Coastal Water CS-5

among natural variations, human-caused pollutant loadings to the Sound, and dissolved oxygen levels in

the Sound. Two models, a water quality model that approximates the biological and chemical processes

of the Sound and a hydrodynamic model that describes physical processes, have been developed. 

An intensive field program in Long Island Sound to collect data for the computer models was undertaken

from April 1988 to September 1989. These data were used to calibrate and verify the models to ensure

that they reproduce the important features of the Sound. The water quality model, called LIS 2.0,

provided needed insight into the causes of hypoxia and was the basis for actions to begin to reduce

nitrogen discharges to the Sound. However, because it simulates the movement of the Sound's waters in

only two dimensions (east-west and surface to bottom) and in a simplified manner, the LIS 2.0 model did

not provide the best technical foundation for identifying the total level of reduction in nitrogen loads that

should be attained or the most cost-effective means to achieve targeted reductions. The hydrodynamic

model, developed by the National Oceanic and Atmospheric Administration and completed in July, 1993,

uses tide and current measurements to simulate the water's circulation in three dimensions (east-west,

north-south, surface to bottom). It was coupled to the water quality model, to create LIS 3.0. The LIS 3.0

model provides an advanced tool to relate sources of nitrogen from specific geographic areas to the

hypoxia problem in the western Sound. Because the impact of the nitrogen load from different

management zones can be deter-mined using LIS 3.0, the LISS can assign priorities for management to

ensure that the most the cost-effective options are pursued.

The modeling, combined with field monitoring and laboratory studies, provided a level of detail to

support some clear conclusions about hypoxia in the Sound, its causes, and its solutions. In addition, the

models allowed the LISS to simulate water quality conditions as they were in the past, as they are today,

and as they could be in the future under alternative nitrogen control scenarios.

• The most oxygen that can be dissolved in Long Island Sound at summer water temperatures is

about 7.5 milligrams per liter (mg/L) of water. This is known as the saturation level. 

• Oxygen concentrations greater than 5.0 mg/L provide healthy conditions for aquatic life. 

Concentrations between 5.0 mg/L and 3.5 mg/L are generally healthy, except for the most sensitive

species. When concentrations fall below 3.5 mg/L, conditions become unhealthy. The most severe

effects occur if concentrations fall below 2.0 mg/L, even for short periods of time.

• The growth of algal blooms in Long Island Sound is dependent upon the availability of nutrients. 

These blooms end when the pool of nitrogen available for continued growth is depleted.

• In pre-colonial days, natural, healthy biological activity brought oxygen levels below saturation due

to the natural loadings of organic material and nitrogen, but oxygen levels probably fell below 5

mg/L only in limited areas and for short periods of time.

• Under today's higher nutrient and organic material loading conditions, minimum oxygen levels

average approximately 1.5 mg/L. These levels are associated with severe hypoxia. 

• By substantially reducing nitrogen loadings to the Sound, the minimum oxygen levels in the bottom

waters during late summer can be increased to an average of about 3.5 mg/L, thereby significantly

reducing the probability and frequency of severe hypoxia and reducing the area affected by

hypoxia.

• Increases in nitrogen delivered to the Sound could significantly worsen the hypoxia problem,

causing larger areas to have lower oxygen levels for longer periods of time. The probability of
CS-6 Nutrient Criteria—Estuarine and Coastal Water

events like the summer of 1987, when anoxia (no oxygen) became a reality in the Sound, offshore

of Hempstead Harbor, would also increase.

Understanding the components of the load of nitrogen entering the Sound is fundamental to

understanding the plan: 

• In 1990, defined as a baseline year by the Management Conference, the total nitrogen load was

90,800 tons per year. 

• By 1992, the total nitrogen load had increased to 93,600 tons per year; this increase was anticipated

and was a consequence of terminating ocean disposal of sewage sludge from New York City and

the need to treat some of the sludge at facilities within the basin, reintroducing nitrogen to the

wastestream. 

• Of the 93,600 tons per year, approximately 39,900 tons are from natural sources and not subject to

reductions by management activity. 

• The remaining 53,700 tons of nitrogen per year are associated with human activities and have the

potential to be reduced through management actions. 

• 10,700 tons of nitrogen per year enter the Sound through its boundaries -- the East River in the west

and The Race in the east; efforts to reduce this substantial western load will come under the

auspices of the New York-New Jersey Harbor Estuary Program. 

• 2,200 tons of nitrogen per year enter the Sound from direct atmospheric deposition; the

Management Conference estimates that this load will be reduced to 1,540 tons of nitrogen per year

through implementation of the 1990 Clean Air Act amendments. 

• The remaining 40,800 tons of nitrogen per year are a result of human activity coming from point

and nonpoint source discharges in the Sound's drainage basin and are the subject of the plan. Point

source discharges, primarily sewage treatment plants, result in 32,400 tons of nitrogen each year

and nonpoint source discharges, such as agricultural and stormwater runoff, result in 8,400 tons of

nitrogen each year. 

The Plan to Solve the Problem

The goal of the hypoxia management plan is to eliminate adverse impacts of hypoxia resulting from

human activities. Achievement of this goal will require very large investments of capital, a long-term

commitment, and the assistance of the New York-New Jersey Harbor Estuary Program. Therefore, the

Management Conference has established interim targets for dissolved oxygen and has outlined a phased

approach to achieving them, using what is known now to support early phases and committing to take

additional steps as increased understanding of the environment will dictate in the future. 

Interim Dissolved Oxygen Targets

Using scientific information on the relationship between oxygen levels and ecological effects, the

Management Conference has established interim target levels for oxygen that, if achieved, would

minimize the adverse impacts of hypoxia. In summary, the interim dissolved oxygen targets for the

bottom waters of the Sound are to: 
Nutrient Criteria—Estuarine and Coastal Water CS-7

• Maintain existing dissolved oxygen levels in waters that currently meet State standards;

• Increase dissolved oxygen levels to meet standards in those areas below the State standards but

above 3.5 mg/L; and,

• Increase short-term average dissolved oxygen levels to 3.5 mg/L in those areas currently below 3.5

mg/L, ensuring that dissolved oxygen never goes below 1.5 mg/L at any time.

• There are also interim targets for the surface waters of the Sound. 

Phased Approach

The Management Conference is implementing a phased approach to reducing nitrogen loadings to the

Sound from point and nonpoint source discharges within the Sound's drainage basin. 

• Phase I, as announced in December of 1990, froze nitrogen loadings to the Sound in critical areas at

1990 levels to prevent hypoxia from worsening. 

• Phase II, as detailed in the plan, includes significant, low-cost nitrogen reductions that begin the

process of reducing the severity and extent of hypoxia in the Sound.

• Phase III will present nitrogen reduction targets to meet the interim targets for dissolved oxygen,

which will prevent most known lethal and sublethal effects of hypoxia on the Sound's estuarine life. 

Phase III also will lay out the approach for meeting the nitrogen load reduction targets. 

Phase I - The Nitrogen Loading Freeze

Phase I was announced in December 1990. It called for a freeze on point and nonpoint nitrogen loadings

to the Sound in critical areas at 1990 levels. It committed the States and local governments to specific

actions to stop a 300-year trend of ever increasing amounts of nitrogen entering the Sound. 

Since 1990, activities have been underway in New York and Connecticut to manage nitrogen from

sources within the New York and Connecticut portions of the drainage basin, starting with adoption of

the Phase I “freeze” on loadings.

• Connecticut reacted quickly to obtain $15 million in State funds to ensure that the nitrogen freeze

was implemented. Consent orders are in place to cap the nitrogen loads at the 15 affected facilities.

• In New York City, the New York State Department of Environmental Conservation (NYSDEC) and

the city have reached full agreement on sewage treatment permit limits, freezing total nitrogen

loadings at 1990 levels. 

• In Westchester County, the NYSDEC has issued final permits to the four existing sewage treatment

plants, freezing their aggregate load at the 1990 level. This was done with the full agreement of the

county. 

• On Long Island, the NYSDEC proposed individual permits that freeze the loads from individual

discharges at 1990 levels; in response, the dischargers proposed establishment of an aggregate

limit. State and local authorities agreed on aggregate load limits for targeted facilities.
CS-8 Nutrient Criteria—Estuarine and Coastal Water

Phase I agreements to control nonpoint sources centered around three categories: 

• Use of existing nonpoint source and stormwater management programs to focus on nitrogen control

with the objective of freezing the loads. 

• Assessing tributary loads to Long Island Sound to begin planning for their control. 

• Assigning priorities for management to coastal subbasins where nitrogen loads were estimated to be

the highest. 

Phase II - Low Cost Nitrogen Reductions

For Phase II, the LISS made a commitment in 1994 to reduce nitrogen discharges to the Sound from peak

loadings by approximately 7,550 tons per year. This phase consists of incorporating a variety of low-cost

nitrogen removal technologies at selected sewage treatment plants. The States have moved aggressively

to implement nitrogen control activities, using innovative strategies and seeking the cooperation of local

governments. 

In Connecticut, the goal was to achieve a reduction of 850 tons per year in nitrogen loads. The State of

Connecticut has awarded more than $15 million through its State Clean Water Fund to 11 southwestern

sewage treatment plants to test and demonstrate the efficiency of upgrades for nitrogen treatment. In

addition, the first plant in the State designed to denitrify has been constructed in Seymour. As of

December 1997, the load of nitrogen from plants in the Phase II agreement has been reduced by almost

900 tons per year, exceeding the Phase II goal. 

The State of New York revised the permits issued to sewage treatment plants, with the consent of local

authorities, to establish nitrogen limits at 1990 levels. The permits include an aggregate load for

facilities within Management Zones 7-11 (New York City, Westchester County, and Long Island). The

New York goal was to reduce nitrogen loadings by 6,700 tons per year from peak loadings from actions

to be completed by 2006. The goal of these actions was to compensate for the increased load due to

sludge treatment and reduce loadings back below 1990 levels. As of 1997, one sewage treatment plant in

Westchester County and four in New York City have implemented nitrogen removal technologies. New

York City is required to implement additional nitrogen removal technologies at the upper East River

sewage treatment plants. As of December 1997, the load of nitrogen from sewage treatment plants in

New York had decreased by 3,000 tons per year from peak loadings. In addition, New York City has

entered into a consent order to provide nitrogen removal at the reconstructed Newtown Creek facility,

scheduled for completion in 2007. 

In addition, both States have: 

• Developed materials and conducted training for treatment plant personnel on nitrogen removal

technologies and procedures. 

• Required sewage treatment plants to identify in their plans how they will remove nitrogen, if

required to do so.

• Required nutrient monitoring at sewage treatment plants to improve understanding of nitrogen

sources and treatment plant capability.

• Increased the share of nonpoint source pollution control funds targeted to projects that reduce

nitrogen loads to the Sound.
Nutrient Criteria—Estuarine and Coastal Water CS-9

• Formulated Coastal Nonpoint Pollution Control Programs to address coastal nonpoint sources of

nitrogen.

• Undertaken demonstration projects that address a variety of nonpoint source control issues and

technologies (e.g., urban runoff treatment by artificial pond/wetland systems, parking lot runoff

treatment, septic system technologies to treat and remove nitrogen, controlling runoff from

agricultural land and from marinas).

As of December 31, 1997, nitrogen loadings to the Sound from point and non-point sources within the

New York and Connecticut portions of the watershed have been reduced as a result of these activities by

3,900 tons per year from peak loadings.

Phase III - Nitrogen Reduction Targets to Eliminate Severe Hypoxia

While steps taken in Phases I and II will help to reduce the extent of hypoxia, additional nitrogen

reduction is needed to restore the health of Long Island Sound. Phase III sets the course by setting

specific nitrogen reduction targets for each of the 11 management zones around the

Sound. An array of environmental and economic considerations were taken into account throughout the

process.

Oxygen Benchmarks

The water quality standard for oxygen in Long Island Sound is 6 mg/L in Connecticut and 5 mg/L in New

York. Modeling indicates that even if maximum nitrogen reduction technologies were implemented, the

water quality standards for oxygen would not be achieved throughout the summer in all areas of the

Sound. To help establish priorities for action, the LISS has identified oxygen conditions that will

minimize adverse impacts on living resources of the Sound. 

Two major research efforts, a laboratory study by the EPA’s Office of Research and Development and a

field study by the Connecticut Department of Environmental Protection (CTDEP) have provided much of

the information on how low oxygen conditions affect living resources in the Sound. Both studies

corroborated that severe effects occurred whenever levels of oxygen fell below 2.0 mg/L. The field

surveys noted large reductions in the number and types of aquatic life present. The lab experiments

recorded reductions in growth and increases in mortality. In both studies, effects became significant

when oxygen levels fell below 3.5 mg/L, though some effects occurred at levels between 3.5-5.0 mg/L. 

As a result, the LISS has determined that unhealthy conditions occur whenever oxygen levels fall below

2.0 mg/L at any-time or remain below 3.5 mg/L over a 24-hour period. Most adverse impacts can be 

prevented if oxygen levels exceed these conditions, and they have been used as benchmarks to assess the

relative benefits of alternative management strategies for improving the health of Long Island Sound.

Cost-effectiveness

LISS managers looked at a range of nitrogen reduction options for the three major sources of nitrogen in

the watershed, sewage treatment plants, industrial facilities, and nonpoint source runoff to determine the

most cost-effective option. 

• Sewage Treatment Plants: As nitrogen removal requirements become more stringent, the cost of

controls tends to increase. To identify a cost-effective level of treatment, LISS managers arrayed

the possible nitrogen reduction options for all 70 sewage treatment plants in the 11 management

zones and calculated the average oxygen improvement in the Sound per dollar spent. 

Improvements at sewage treatment plants that had better than average cost-effectiveness at

improving oxygen conditions in the Sound were identified. These actions, in total, could achieve a

62 percent reduction in loads, or 122,044 pounds/day.
CS-10 Nutrient Criteria—Estuarine and Coastal Water

• Industrial Facilities: A limited number of industrial facilities directly contribute nitrogen to the

Sound; all are located in Connecticut and contribute an estimated 6,717 pounds per day of nitrogen

to the Sound. Because information on the cost of reducing nitrogen from industrial sources was not

readily available, these facilities were not included in the cost analyses used for sewage treatment

plants. Instead, the cost-effective level of treatment identified for sewage treatment plants, 62

percent, was applied to the industrial sources, resulting in a 4,165 pounds per day reduction for

industrial facilities. This represents an aggressive but cost-effective level of nitrogen control for

these sources.

• Nonpoint Sources: Decisions on controls of nonpoint source runoff must be made in the broader

context of watershed management, since control measures will also help reduce suspended solids,

toxic contaminants, pathogens, and floatable debris. The LISS recommends that aggressive controls

of nonpoint source pollution be implemented for both existing and new development, through both

habitat protection and restoration activities, and structural and nonstructural best management

practices. This effort could result in a 10 percent reduction in the non-point source load from

sources within the New York and Connecticut portions of the watershed, or 2,604 pounds per day. 

Adding the potential nitrogen reductions from cost-effective controls on sewage treatment plants,

industrial sources, and nonpoint runoff sources results in a total reduction of 128,813 pounds per day

(23,500 tons per year). The next step is to allocate responsibility for achieving these reductions among

the 11 management zones fairly. 

Allocating Responsibility

The cost curve analysis provided an option for allocating nitrogen reductions among the sewage

treatment plants. Sewage treatment plant upgrades with greater than average cost-effectiveness would be

implemented while upgrades with below average cost-effectiveness would not be implemented. 

However, the LISS decided that relying on the cost curve analysis alone would not be a fair or even

feasible approach and would not provide the best solution to allocating nitrogen reduction. 

There are several reasons for this conclusion. Most importantly, the cost estimates were general and not

uniform in their development. More accurate cost estimates must await detailed facilities planning based

upon a clear definition of the nitrogen discharge limits that will have to be met. In addition, local

concerns and considerations such as the need to purchase land for expansion and to distinguish between

costs for nitrogen removal versus ongoing maintenance, expansions for growth, and secondary upgrade

needs (which were not included in the cost estimates) were not addressed evenly in the cost analysis. 

Cost considerations aside, it is necessary for all sewage treatment plants to share the burden of nitrogen

removal. All sewage treatment plants contribute nitrogen to Long Island Sound, albeit with different

effect. All jurisdictions will benefit from improved water quality. Therefore, it is reasonable to expect

all contributors to the problem to con-tribute to the solution.

For those reasons, LISS has assigned each management zone equal responsibility to reduce its share of

the nitrogen load. To achieve a similar level of oxygen improvement from reductions allocated to each

zone by the same percentage, the load reduction target was adjusted slightly to 23,800 tons per year from

the original 23,500 tons per year. The total human-derived load coming from sewage treatment plants,

industrial point sources, and nonpoint sources, including atmospheric depositions within the water-shed,

is 40,650 tons per year. Therefore, the Soundwide nitrogen target is a 58.5 percent reduction in the

human-derived load from point and nonpoint sources in the watershed.
Nutrient Criteria—Estuarine and Coastal Water CS-11

Phase III Actions

Phase III actions will minimize adverse impacts of hypoxia caused by human activities in

a cost-effective manner, while ensuring that new information is gathered to refine and improve

management over the long term. Using the framework described above, the LISS set a 58.5 percent

reduction target for the enriched load of nitrogen from sources within the New York and Connecticut

portions of the watershed.

Strategies

Attaining the nitrogen reduction targets will require aggressive control of point sources, such as sewage

treatment plants and industrial sources, and nonpoint sources, such as on-site sewage systems and runoff

from roads, parking lots, and construction sites. To achieve the reduction targets, the States, working

with local governments, will select the mix of point and nonpoint source controls to be implemented in

each management zone. Recognizing that each watershed is different, the plan provides the States and

municipalities considerable flexibility in determining how nitrogen reduction actions are carried out

within each zone. 

By August 2000, the States will take the following actions:

• Develop watershed plans for each management zone that will set the course for achieving the

targets as scheduled. 

• Consistent with those plans, incorporate limits on the amount of nitrogen that can be discharged

from sewage treatment plants and industrial sources into discharge permits. 

• Conduct comprehensive nonpoint source management and habitat restoration activities. 

Because the total nitrogen load entering the Sound from human sources is dominated by point source

discharges, the plan emphasizes technologies that can be applied to sewage treatment facilities and

industrial discharges.

In order to achieve significant reductions in the nonpoint source nitrogen load, home owners, farmers,

businesses, municipalities, and the States will need to reduce current inputs of nitrogen to the watershed

and restore and preserve the nitrogen removal capabilities of existing natural systems. These reductions

can be achieved using a number of approaches—resource-based land use decisions at the local level,

watershed-wide use of appropriate structural and nonstructural best management practices (e.g.,

stormwater detention ponds, artificial wetlands, streetsweeping, cleaning catch basins), habitat protection

and restoration, and pollution prevention management practices. All approaches will require a concerted

education and outreach effort. 

Timing

The planning, financing, and construction of upgrades to sewage treatment plants necessary to achieve

the 58.5 percent reduction target will require sustained effort and commitment over a long period of time. 

Therefore, the LISS recommends phasing-in the necessary reductions over 15 years: 

• 40 percent in 5 years,

• 75 percent in 10 years, and

• 100 percent in 15 years.
CS-12 Nutrient Criteria—Estuarine and Coastal Water

Cost

The Comprehensive Conservation and Management Plan identified that the cost of achieving maximum

nitrogen removal from all point sources would range from $6 to $8 billion ($5.1 to $6.4 billion in New

York State and from $900 million to $1.7 billion in Connecticut). Because of the successful

demonstration of full scale nitrogen removal technologies at sewage treatment plants undertaken as part

of Phase II, the estimated costs of capital improvements at sewage treatment plants have decreased. The

estimated cost of achieving maximum nitrogen removal levels at the 70 treatment plants in New York

and Connecticut is now about $2.5 billion 

Because of the cost-effective approach described above, the LISS nitrogen reduction strategy would not

require all treatment plants to meet limit-of-technology reductions. As a result, the incremental capital

cost of achieving the Phase III point source controls was estimated to be $300 million for New York State

and $350 million for Connecticut. These cost estimates have been questioned and will be revised as

more detailed facility planning and design is performed. However, they show clearly that the potential

cost of achieving our goals can be much less than originally estimated. 

Nonpoint source controls will be implemented as part of broader watershed and habitat protection efforts. 

The cost of controlling nonpoint sources is more difficult to estimate than the cost of point source

controls. Rather than one type of technology applied to a similar source, a variety of strategies can be

applied to control a variety of nonpoint sources of nitrogen. As a result, the costs of achieving nonpoint

nitrogen reductions will be addressed in the zone-by-zone plans developed by the States.

Financing

As recommended in the Comprehensive Conservation and Management Plan, the main source of funding

for these wastewater treatment facility improvements will be the State Revolving Fund programs. The

EPA, through the federal Clean Water Act, provides financing to support State Revolving Fund loan

programs. 

Connecticut uses the capitalization grant from EPA to leverage with State bond funds to provide grants

and low interest loans, at 2 percent interest over 20 years, to finance improvements at municipal

facilities. Connecticut provides about $50 million per year in State bonding to supplement the $15

million per year provided under the Clean Water Act. At this capitalization rate, Connecticut should be

able to meet municipal financing needs to implement Phase III nitrogen reductions. During fiscal year

1997, CTDEP awarded $250 million from their Clean Water Fund to finance projects of benefit to Long

Island Sound, including major sewage treatment plant upgrades in Norwalk and Waterbury.

New York State established its State Revolving Fund in the custody of the Environmental Facilities

Corporation. This public corporation benefits local governments in New York State by offering belowmarket interest rate loans to municipalities to finance wastewater improvements. Currently, the interest

rate is set at up to one-half of the market rate to be repaid in 20 years. Lower rates of interest, including

zero interest loans, are available for communities that can demonstrate an inability to pay the standard

subsidized rate. Another major source of funding in New York State is the $1.75 billion Clean

Water/Clean Air Bond Act approved by voters in November 1996. The Bond Act targeted $200 million

for Long Island Sound that will be available for sewage treatment upgrades, habitat restoration, nonpoint

source control, and pollution prevention.

The possible funding sources for non-point source controls reflect the diversity of both the sources and

the control options. Grant funding through federal and State water quality management, natural

resources management, and coastal zone management programs is available for nonpoint source

activities. The State Revolving Fund loan program is also available to fund stormwater management and
Nutrient Criteria—Estuarine and Coastal Water CS-13

habitat restoration projects but has not been used to a great extent for these types of activities due to the

magnitude of existing point source funding needs in Connecticut and New York.

Effluent Trading

To provide further flexibility and incentives for maximizing the timeliness and cost-effectiveness of

nitrogen reduction actions, the LISS is investigating the feasibility of allowing effluent trading. Trading,

if employed as part of the nitrogen reduction effort, may be an innovative way to use market forces to

more efficiently meet water quality goals. The LISS is developing a trading proposal and will convene a

public forum for federal, State, and local water quality officials, together with public and private

interests, to evaluate its potential.

Enforcement

The provisions of the federal Clean Water Act provide a vehicle for ensuring that nitrogen reduction

targets are legally enforceable. Section 303(d) of the Act requires the identification of a Total Maximum

Daily Load for pollutants that will result in the attainment of water quality standards. Once a Total

Maximum Daily Load has been established, the act calls for reductions to be allocated to sources so that

the load target is met. New York and Connecticut and EPA will use their authorities to provide an

enforceable foundation for achieving the nitrogen reduction targets. By August, 1998 the States will

propose a Total Maximum Daily Load designed to meet State oxygen standards. The current Long Island

Sound standards were developed with limited data on how low oxygen levels affect aquatic life in Long

Island Sound. EPA is currently developing regional marine oxygen criteria that will provide a more

scientifically valid basis for the development of oxygen standards. Based on this information, the States

may, in the future, modify their oxygen standards. While LISS managers predict significant

improvement in water quality as the nitrogen reduction targets are implemented, the attainment of current

water quality standards at all times and in all areas is not expected. For this reason, the LISS will

continue to assess what other kinds of actions will be needed to bring the Sound into full compliance

with water quality standards.

These actions may include control of nitrogen and carbon sources outside of the Long Island Sound basin

(e.g., tributary import from point and non-point sources north of Connecticut, atmospheric deposition,

boundary import from point and nonpoint sources affecting New York Harbor and The Race). 

Alternatives to nitrogen reduction, such as aeration, will need to be considered as a possible means to

achieve water quality standards in remaining areas.

Evaluating Progress

The LISS will track, monitor, and report on progress in meeting the nitrogen reduction targets annually.

In addition, a formal review of the goals and objectives of the program will be performed every 5 years,

coinciding with the progress checkpoints for nitrogen reduction. The review will consider: 

• Progress and cost of implementation, including a reevaluation of the knee-of-the-curve analysis

used to establish the Phase III nitrogen reduction targets,

• Improvements in technology, including the results of quality controlled pilot projects,

• The regional dissolved oxygen criteria to be published for comment, 

• Water quality standards, 

• Refined information on the ecosystem response to nitrogen reductions,
CS-14 Nutrient Criteria—Estuarine and Coastal Water

• The results of peer reviewed modeling, and

• Research on the impacts of hypoxia to living resources and their habitats.

Each of these factors will be considered in a balanced manner in the reevaluation process. As a result of 

the review, the LISS may recommend improvements that could result in changes in how the overall

program will be implemented.

For More Information:

Mark Tedesco

EPA Long Island Sound Office

888 Washington Blvd.

Stamford, CT 06904-2152

Phone: 203/ 977-1541

Fax: 203/ 977-1546
Nutrient Criteria—Estuarine and Coastal Water CS-15

CASE STUDY

NP BUDGET FOR NARRAGANSETT BAY

S. V. Smith

University of Hawaii 

808-956-8693

http://data.ecology.su.se/MNODE/North%20America/NRB.HTM 

Narragansett Bay, Rhode Island (41/ 35’ N, 71/ 20’W), is a relatively well mixed, near-oceanic salinity

estuary on the northeast (Atlantic) coast of the U. S. It occupies an area of 264 km2 (Table 1) and has a

mean depth of 9.7 m. Note that both the area and volume differ from the comments in Nixon et al.

(1995), but seem consistent with Kremer and Nixons (1978) explicit tabulation. Freshwater flow into the

system averages about 8.2 x 106 m

3

/d, from a watershed of 3500 km2. Primary production in the system is

dominated by phytoplankton (29 mol C m-2 yr-1) with a C:N:P ratio of about 112:13:1. The budget

described below is based on data collected primarily in the late 1970’s and through much of the 1980’s. 

Details of this kind of analysis can be found at the LOICZ - Biogeochemical Modelling web site at: 

http://data.ecology.su.se/MNODE/index.htm.

Sector area and volume data are from Kremer and Nixon (1978). Sector nutrient concentrations are

annual averages (based on surface and deep water data) also from Kremer and Nixon. Sector nutrient

masses are calculated as volume x concentration. The sectors at the bay mouth (#5, 8) are used for

"oceanic values."

Nutrient exchange fluxes (Table 2) are calculated using an average 26-day exchange time, as calculated

by Pilson (1985) with a water and salt budget (analogous to procedure in Gordon et al., 1996). The bay 

Table 1. Sector areas, volumes, and nutrient concentrations. Data are used to calculate volumeaveraged concentrations for the outer portion of the bay ("ocean") and the bay proper

SECT.

#

VOL.106

m3

AREA

106 m2

DIP

:M

NH4

:M

NO3

:M

Sum DIN

:M

DIP

106 mol

NH4

106 mol

NO3

106 mol

S DIN

106 mol

1 130 20.1 1.8 12 11 23 0.23 1.56 1.43 2.99 

2 300 44.6 1.5 7 6 13 0.45 2.10 1.80 3.90 

3 115 28.5 1.6 3 6 9 0.18 0.35 0.70 1.05 

4 463 61.9 1.4 4 4 8 0.65 1.85 1.85 3.70 

5 204 20.0 1.0 1 3 4 0.20 0.27 0.60 0.87 

6 222 26.0 1.2 2 5 7 0.27 0.44 1.11 1.55 

7 573 38.5 1.0 2 4 6 0.57 1.15 2.29 3.44 

8 554 24.2 0.7 1 2 3 0.39 0.39 1.11 1.50 

SUM 2561 264 

bay 1.3 4 5 9 

ocean 

(secs.

# 5,8)

0.8 1 2 3 
CS-16 Nutrient Criteria—Estuarine and Coastal Water

Table 2. Hydrographic exchange fluxes of nutrients

SOURCE OF 

FLUX

DIP 

106mol/yr 

NH4

106mol/yr 

NO3

106mol/yr 

DIN

106mol/yr 

residual flow -3 -7 -11 -18 

net exchange flow -18 -108 -108 -216 

total hydrography -21 -115 -119 -234 

volume divided by the residence time gives a mixing exchange volume of 98.5 x 106 m3/d, while the

residual outflow equals the freshwater inflow (i.e., 8.2 x 106 m

3

/d). It would, in principle, be possible to

time-step through the data (at monthly increments, for example). However, to do that would require

having the flow data to go with the nutrient data. Further, inspection of the graph by Nixon et al. (1995)

of flow data and comparison of Pilson’s (1985) flow—residence time regression equation suggests that

the residence time over this range of flow is well approximated by a constant value for the exchange

time. Various authors describe the bay as well mixed, and this is supported by the water composition data

in Kremer and Nixon (1978). We therefore use a 1-box model to perform these calculations, rather than

a vertically stratified model to describe hydrographic fluxes.

In Table 3, all boundary fluxes except hydrography were taken directly from by Nixon et al. (1985).

Hydrographic flux was calculated as above. ) Y’s (the nonconservative fluxes) are calculated by

difference (as described in Gordon et al., 1996). No data are available for DOP, DON, or for either

inorganic or organic C, so the budget is based on inorganic N and P only. As discussed by Gordon et al. 

and consistent with comments in Nixon et al., it seems safe to assume that DOP and DON

nonconservative fluxes do not contribute strongly to the overall nonconservative fluxes in this system.

Rates for the D Y’s per unit area are calculated using the bay area of 264 km2 (Table 4). Note that this

area estimate is about 25% lower than the value used by Kremer and Nixon (1978). We have used the

smaller area and volume on the basis that these are the data used to calculate the volume-averaged

concentrations. Net (nfix-denit) is calculated on the assumption that the N:P ratio of D DIP is 13:1, then

D DIN is balanced. Net (p-r) is calculated from the DIP flux, using a C:P ratio of 112:1.

Table 3. Total boundary fluxes of nutrients and inferred internal reactions—

the system budget

Process

DIP 

106mol/yr 

NH4

106mol/yr 

NO3

106mol/yr 

DIN

106mol/yr 

atmosphere 0 6 19 25 

rivers 13 113 177 290 

urban runoff 2 13 4 17 

sewage 9 136 6 142 

hydrography -21 -115 -119 -234 

D Y -3 -153 -87 -240 

(nfix-denit) -201 
Nutrient Criteria—Estuarine and Coastal Water CS-17

Table 4. Nonconservative fluxes of materials and stoichiometrically inferred

biogeochemical pathways

DIP

mmol m-2 yr-1

NH4

mmol m-2 yr-1

NO3

mmol m-2 yr-1

DIN

mmol m-2 yr-1

C

mol m-2 yr-1

D Y -11 -580 -329 -909 

D DINexp -143 

(nfix-denit) -766 

(p-r) 1.2 

Nixon et al. (1995) have data with which the present budgetary estimates may be compared: They

estimate DIP and DIN fluxes from the ocean to the bay by a hydrographic budget analogous to values

used here for both influx and efflux, but they do not use this same hydrography to estimate nutrient

fluxes to the ocean. Their inward DIP and DIN fluxes, obtained by time-stepping through the oceanic

nutrient concentration data (bottom water only), are 27 and 115 x 106 mol/yr. The calculations here

(using annual average data) are 29 and 108 x 106 mol/yr. The agreement is within 10%. It should be

close, because both Nixon et al. and the calculations here are performing essentially the same calculation. 

Three points for minor disagreement would be that the values here just used a constant exchange rate

(instead of time-varying); values used here were picked data off a graph; and surface and bottom values

were averaged (on the graph, these are effectively identical in the outermost bay sectors). 

Instead of using hydrography to estimate outward DIP and DIN flux, those authors estimate DIP and DIN

fluxes from the bay to ocean by difference with other terms in their budget, to close the budget. They get

41-51 x 106 and 240-470 x 106 mol/yr. Again pulling the hydrographic terms apart, the calculations here

yield 50 x 106 and 342 x 106 mol/yr (in both cases, within their range). It is worth noting that if the water

exchange volume is incorrect, it would affect both influx and efflux of nutrients, hence have a relatively

small effect on the difference between influx and efflux. The point here, of course, is that the difference

between influx and efflux is probably more reliable than either of the individual fluxes.

Nixon et al. use a variety of considerations for two different sets of incubation data to assign baywide

denitrification a range of 85-170 x 106 mol/yr (compared to 201 x 106 from the hydrographic budget;

using their high values, agreement is within 20%). 

Those authors estimate respiration to consume 8100 to 9200 x 106 mol/yr of organic C. Using their

estimate for primary production (p) of 29 mol C m-2 yr-1 and the DIP-derived estimate for production -

respiration (p-r) of 1.2, r is estimated to be 27.8 mol C m-2 yr-1. Scaling by the bay area, this gives

respiration to be 7340 x 106

 mol/yr (within 20% of their lower estimate). If we were to use the are value

given in Nixon et al. (328 km2, instead of the value of 264 km2 from Kremer and Nixon (1978), the

respiration would be 9118 x 106 mol/yr (within their range).

Efforts to control the release of nutrients into Narragansett Bay have recently addressed nitrogen

contributions from Publically Owned Treatment Works (POTWs) throughout the watershed. One

nutrient reduction option currently being pursued is to maximize nitrogen removal from the final effluent

by modifying operating conditions with existing equipment at the facility. Retrofitting existing facilities

will also be considered where appropriate. A second venue involves drafting water quality based permit

limits over the next few years to limit nitrogen in the final effluent of POTWs. Finally, a total maximum

daily load (TMDL) for nitrogen is currently under development for the Providence River upstream of

Narragansett Bay through the NPDES permitting process. A model is being developed that once

calibrated, will set nitrogen load limits for POTWs that discharge to the river.
CS-18 Nutrient Criteria—Estuarine and Coastal Water

For Further Information:

S. V. Smith 

Department of Oceanography, 

University of Hawaii 

1000 Pope Road Honolulu, Hawaii 96822 USA 

email: svsmith@soest.hawaii.edu 

phone: 808-956-8693 

fax: 808-956-7112
Nutrient Criteria—Estuarine and Coastal Water CS-19

CASE STUDY

NUTRIENT MANAGEMENT AND SEAGRASS RESTORATION IN TAMPA BAY, FLORIDA

Holly Greening, Tampa Bay Estuary Program

Abstract: Participants in the Tampa Bay Estuary Program have agreed to adopt nitrogen loading

targets for Tampa Bay based on the water quality and related light requirements of the seagrass

species Thalassia testudinum. Based on modeling results, it appears that light levels can be

maintained at necessary levels by “holding the line” at existing nitrogen loadings. However, this

goal may be difficult to achieve given the 20% increase in the watershed’s human population and

associated 7% increase in nitrogen loading that are projected to occur over the next 10-20 years.

To address the long-term management of nitrogen sources, a Nitrogen Management Consortium of 

local electric utilities, industries and agricultural interests, as well as local governments and

regulatory agency representatives, has developed a Consortium Action Plan to address the target

load reduction needed to “hold the line” at 1992-1994 levels. To date, implemented and planned

projects collated in the Consortium Action Plan meet and exceed the agreed-upon nitrogen loading

reduction goal.

The Tampa Bay estuary is located on the eastern shore of the Gulf of Mexico in Florida, USA. At more

than 1000 km2, it is Florida’s largest open water estuary. More than 2 million people live in the 5700

km2 watershed, with a 20% increase in population projected by 2010. Land use in the watershed is

mixed, with about 40% of the watershed undeveloped, 35% agricultural, 16% residential, and the

remaining commercial and mining.

Major habitats in the Tampa Bay estuary include mangroves, salt marshes and submerged aquatic

vegetation. Each of these habitats has experienced significant areal reductions since the 1950s, due to

physical disturbance (dredge and fill operations) and water quality degradation, particularly impacting

the seagrasses due to loss of light availability. Five species of seagrass are commonly found in Tampa

Bay, with Thalassia testudinum (turtlegrass) and Syringodium filiforme (manatee grass) dominating in

the higher salinity areas and Halodule wrightii (shoalgrass) and Ruppia maritima (widgeon grass) most

commonly found in lower salinities. 

The importance of seagrass as a critical habitat and nursery area for fish and invertebrates, and as a food

resource for manatees, sea turtles and other estuarine organisms has been recognized by the Tampa Bay

resource management community for several decades. In 1990, Tampa Bay was accepted into the U.S.

Environmental Protection Agency’s (EPA) National Estuary Program. The Tampa Bay National Estuary

Program (TBNEP), a partnership that includes three regulatory agencies and six local governments, has

built on the resource-based approach initiated by earlier bay management efforts. Further, it has

developed water quality models to quantify linkages between nitrogen loadings and bay water quality,

and models that link water quality to seagrass goals.

Recent recommendations from the National Academy of Science National Research Council (NRC)

include those which regional watershed programs might consider in developing nutrient management

strategies. The NRC recommendations are based on the process designed by the Tampa Bay Estuary

Program partners to develop and implement a seagrass protection and restoration management program

for Tampa Bay. Critical elements of the Tampa Bay process are to:
CS-20 Nutrient Criteria—Estuarine and Coastal Water

1. Set specific, quantitative seagrass coverage goals for each bay segment.

2. Determine seagrass water quality requirements and appropriate nitrogen loading targets.

3. Define and implement nitrogen management strategies needed to achieve load management targets.

STEP 1. SET QUANTITATIVE RESOURCE MANAGEMENT GOALS

Establishment of clearly defined and measurable goals is crucial for a successful resource management

effort. In 1992, TBNEP adopted an initial goal to increase current Tampa Bay seagrass cover to 95% of

that present in 1950.

Based on digitized aerial photographic images, it was estimated that approximately 16,500 ha of seagrass

existed in Tampa Bay in 1950. At that time, seagrasses grew to depths of 1.5 m to 2 m in most areas of

the bay. By 1992, approximately 10,400 ha of seagrass remained in Tampa Bay, a loss of more than 35%

since the 1950 benchmark period. Some (about 160 ha) of the observed loss occurred as the result of

direct habitat destruction associated with the construction of navigation channels and other dredging and

filling projects within existing seagrass meadows, and is assumed to be nonrestorable through water

quality management actions.

In 1996, the TBNEP adopted a bay-wide minimum seagrass goal of 15,400 ha. This goal represented

95% of the estimated 1950 seagrass cover (minus the nonrestorable areas), and includes the protection of

the existing 10,400 ha plus the restoration of an additional 5,000 ha.

STEP 2. DETERMINE SEAGRASS WATER QUALITY REQUIREMENTS AND APPROPRIATE

NITROGEN LOADING RATES

Once seagrass restoration and protection goals were established by the participants, the next steps

established the environmental requirements necessary to meet agreed-upon goals and subsequent

management actions necessary to meet those requirements.

A. Determine environmental requirements needed to meet the seagrass restoration goal

Recent research indicates that the deep edges of Thalassia testudinum meadows, the primary seagrass

species for which nitrogen loading targets are being set, correspond to the depth at which 20.5% of

subsurface irradiance (the light that penetrates the water surface) reaches the bay bottom on an annual

average basis. The long-term seagrass coverage goal can thus be restated as a water clarity and light

penetration target. Therefore, in order to restore seagrass to near 1950 levels in a given bay segment,

water clarity in that segment should be restored to the point that allows 20.5% of subsurface irradiance to

reach the same depths that were reached in 1950.

B. Determine water clarity necessary to allow adequate light to penetrate to the 1950 seagrass deep edges

Water clarity and light penetration in Tampa Bay are affected by a number of factors, such as

phytoplankton biomass, non-phytoplankton turbidity, and water color. Water color may be an important

cause of light attenuation in some bay segments; however, including color in the regression model did not

produce a significant improvement in the predictive ability of the regression model. Results of the

modeling effort indicate that, on a baywide basis, variation in chlorophyll a concentration is the major

factor affecting variation in average annual water clarity.

C. Determine chlorophyll a concentration targets necessary to maintain water clarity needed to meet the

seagrass light requirement

An empirical regression model was used to estimate chlorophyll a concentrations necessary to maintain

water clarity needed for seagrass growth for each major bay segment. The adopted segment-specific
Nutrient Criteria—Estuarine and Coastal Water CS-21

annual average chlorophyll a targets (ranging from 4.6 :g/l to 13.2 :g/l) are easily measured and tracked

through time, and are used as intermediate measures for assessing success in maintaining water quality

requirements necessary to meet the long-term seagrass goal. 

D. Determine nutrient loadings necessary to achieve and maintain the chlorophyll a targets

Water quality conditions in 1992-1994 appear to allow an annual average of more than 20.5% of

subsurface irradiance to reach target depths (i.e., the depths to which seagrasses grew in 1950) in three of

the four largest bay segments. Thus, a management strategy based on “holding the line” at 1992-1994

nitrogen loading rates should be adequate to achieve the seagrass restoration goals in these segments. 

This “hold the line” approach, combined with careful monitoring of water quality and seagrass extent,

was adopted by the TBNEP partnership in 1996 as its initial nitrogen load management strategy.

As an additional complicating factor, a successful adherence to the “hold the line” nitrogen loading

strategy may be hindered by the projected population growth in the watershed. A 20% increase in

population, and a 7% increase in annual nitrogen load, are anticipated by the year 2010. Therefore, if the

projected loading increase (a total of 17 U.S. tons per year) is not prevented or precluded by watershed

management actions, the “hold the line” load management strategy will not be achieved.

STEP 3. DEFINE AND IMPLEMENT NITROGEN MANAGEMENT STRATEGIES NEEDED TO

ACHIEVE LOAD MANAGEMENT GOALS

Local government and agency partners in the TBNEP signed an Intergovernmental Agreement (IA) in

1998 pledging to carry out specific actions needed to “hold the line” on nitrogen loadings. The IA

includes the responsibility of each partner for meeting the nitrogen management goals, and a timetable

for achieving them. How those goals are reached will be left up to the individual communities as defined

by them in their Action Plans. The Tampa Bay National Estuary Program was also renamed the Tampa

Bay Estuary Program as part of the progression from the planning phase to implementation of the

adopted Comprehensive Conservation and Management Plan. 

To maintain nitrogen loadings at 1992-1994 levels, local government Action Plans address that portion of

the nitrogen target which relates to non-agricultural stormwater runoff and municipal point sources

within their jurisdictions, a total of 6 U.S. tons of nitrogen per year through the year 2010 (Table 1). 

To address the remaining 11 U.S. tons of nitrogen of the 17 total per year each year through the year

2010 needed to “hold the line” (attributed to atmospheric deposition, industrial and agricultural sources

and springs), a Nitrogen Management Consortium of local electric utilities, industries and agricultural

interests, as well as the local governments and regulatory agency representatives in the TBEP, was

established (Table 2). The Nitrogen Management Action Plan developed by public and private partners

in the Consortium combines for each bay segment all local government, agency and industry projects that

will contribute to meeting the five year nitrogen management goal. To ensure that each partner was

using similar nitrogen load reduction assumptions for similar projects, guidelines for calculating nitrogen

load reduction credits were developed with the partners, and were used by each of the partners in the

development of their action plans.

The types of nutrient reduction projects included in the Consortium’s Nitrogen Management Action Plan

range from traditional nutrient reduction projects such as stormwater upgrades, industrial retrofits and

agricultural best management practices to actions not primarily associated with nutrient reduction, such

as land acquisition and habitat restoration projects. A total of 105 projects submitted by local

governments, agencies and industries are included in the Plan; 95% of these projects address nonpoint 
Nutrient Criteria—Estuarine and Coastal Waters CS-22

Table 1. Tampa Bay Nitrogen Management Goals

SOURCE

CATEGORY

CUMULATIVE 1995-1999 GOALS FOR NITROGEN REDUCTION/MANAGEMENT

TOTAL

(reduction in annual load)

(tons) Pinellas

County

City of

Clearwater

City of St.

Petersburg

Hillsborough

County

City of

Tampa

Manatee

County

TB 

Consortium*

Old

Tampa Bay

0.30 0.20 0.05 0.40 0.10 <0.01 1.05 2.10

Hillsborough

Bay

<0.01 <0.01 <0.01 4.75 8.45 <0.01 28.25 41.50

Middle

Tampa Bay

<0.01 <0.01 0.90 2.50 <0.01 0.50 7.15 11.05

Lower

Tampa Bay

<0.01 <0.01 <0.01 <0.01 <0.01 8.35 17.00 25.35

Boca Ciega

Bay

0.85 <0.01 1.05 <0.01 <0.01 <0.01 2.00 3.90

TOTAL 1.15 0.20 2.00 7.65 8.55 8.85 55.45 83.85

% 1.4 0.2 2.4 9.1 10.2 10.6 66.1 100.0

* Tampa Bay Nitrogen Management Consortium
Nutrient Criteria—Estuarine and Coastal Water CS-23

Table 2. Public and Private Partners of the Tampa Bay Nitrogen Management

Consortium, July 2001

Public Partners:

City of Tampa

City of Clearwater

City of St. Petersburg

Manatee County

Hillsborough County

Pinellas County

Manatee County Agricultural Extension Service

Environmental Protection Commission of

Hillsborough County

Tampa Bay Regional Planning Council

Florida Department of Environmental Protection

Florida Fish and Wildlife Commission/Florida

Marine Research Institute

Southwest Florida Water Management District

U.S. Army Corps of Engineers

U.S. Environmental Protection Agency

Tampa Port Authority

Tampa Bay Estuary Program

Florida Department of Agriculture and Consumer

Services

Private Partners:

Florida Phosphate Council

Florida Power & Light

Tampa Electric Company

Florida Strawberry Growers Association

IMC-Phosphate

Cargill Fertilizer, Inc.

CF Industries, Inc.

Pakhoad Dry Bulk Terminals

Eastern Associated Terminals Company

CSX Transportation

sources and account for 71% of the expected total nitrogen reduction. Half (50%) of the total load

reduction will be achieved through public sector projects, and 50% by industry.

Table 3 summarizes expected reductions from those projects which were completed by the end of 1999. 

A total of 134 tons per year reduction in nitrogen loading to Tampa Bay is expected from the completed

projects, which exceeds the 1995-1999 reduction goal of 84 tons per year by 60%. An updated estimate

of nitrogen loadings to the bay from all sources was initiated by TBEP in summer 2001, after which the

effectiveness of the proposed projects in maintaining loads to the bay will be evaluated. 

Examples of specific projects and expected nitrogen loading reductions include the following: 

Stormwater facilities and upgrades: Stormwater improvements or new facilities include both

public and private examples. Stormwater retrofits using alum injection to urban lakes reduced total

nitrogen (TN) loading by an estimated 6.4 tons per year. Stormwater improvements eliminated an

estimated 2 tons of TN loading per year. Industrial stormwater improvements at phosphate 
CS-24 Nutrient Criteria—Estuarine and Coastal Water

Table 3. Tampa Bay Nitrogen Management Consortium Summary of Goals and Expected

Reductions (cumulative tons TN reduced or precluded/year by the year 2000

Bay Segment

1995-1999 Nitrogen 

Reduction Goal

Expected Reduction: Completed 

or Ongoing Projects1

Expected Reduction:

Atmospheric Deposition2

Old Tampa Bay 2.10 5.1 3.6 - 6.2

Hillsborough Bay 41.5 65.9 13.9 - 24.0

Middle Tampa

Bay

11.1 21.1 4.6 - 7.9

Lower Tampa

Bay

25.4 36.2 5.7 - 10.0

Boca Ciega Bay 3.9 5.6 1.2 - 2.1

Total 84.0 133.9 29.0- 50.2

1

 Projects have been completed or are under construction. These summaries do not include reductions expected from atmospheric

deposition reductions.

2

 Range of atmospheric deposition reductions expected, based on two methods.

fertilizer factories and transport terminals reduced almost 20 tons TN loading per year by the year

2000. 

Land acquisition and protection: Land acquisition and maintenance of natural or low intensity

land uses precludes higher density uses and higher rates of TN loading. Land acquisition precluded

more than 15 tons TN loading per year by the end of 1999. 

Approved overlay districts requiring additional nutrient control in management areas precluded an

estimated 10 tons per year TN loading.

Wastewater reuse: Wastewater reuse programs resulted in a 6.4 ton per year reduction on TN

loading. Conversion of septic systems to sewer reduced TN loading by 1.7 tons per year.

Emissions reduction: Estimated emissions reduction from coal-fired electric generating plants

between 1995-1997 resulted in reductions of NOx emissions of 11,700 - 20,000 tons. To estimate

the reduction of nitrogen deposition which reaches the bay (either by direct deposition to the bay’s

surface, or by deposition and transport through the watershed), a 400:1 ratio (NOx emissions units

to nitrogen units entering the bay) is assumed. Expected reductions from atmospheric deposition

thus ranged from 29 to50 tons per year by 1999. To date, emissions reductions have not been

included in the estimated total TN reduction to the bay, pending agreement on estimation methods. 

Habitat restoration: Although typically conducted for reasons other than nutrient reduction,

habitat restoration to natural land uses reduces the amount of TN loading per acre in runoff. 

Habitat restoration projects have been completed or are underway in all segments of Tampa Bay’s

watershed. Estimated TN load reduction from completed habitat restoration projects totaled an

estimated 7 tons per year.
Nutrient Criteria—Estuarine and Coastal Water CS-25

Agricultural BMPs: Water use restrictions have promoted the use of microjet or drip irrigation on

row crops (including winter vegetables and strawberries) and in citrus groves. Micro-irrigation has

resulted in potential water savings of approximately 40% or more over conventional systems and an

estimated 25% decrease in fertilizer applied. Nitrogen reduction estimates from these actions total

6.4 TN tons per year.

Education/public involvement: For those projects for which nitrogen load reductions have not

been calculated or measured, but some reductions are expected, the Consortium Action Plan

assumes a 10% reduction estimate until more definitive information is available. These programs

have reduced TN loading by an estimated 2 tons per year.

Industrial upgrades: A phosphate fertilizer mining and manufacturing plant has terminated the use

of ammonia in flot-plants (an element of the fertilizer manufacturing process), resulting in a

reduction of 21 tons per year of nitrogen loading. Other fertilizer manufacturing companies have

upgraded their product conveyor systems, resulting in a TN reduction of more than an estimated 10

tons per year due to control of fertilizer product loss. The termination of discharge by an orange

juice manufacturing plant into a tributary of Tampa Bay has resulted in a reduction of more than 11

tons per year TN loading.

The approach advocated by the TBEP stresses cooperative solutions and flexible strategies to meet

nitrogen management goals. This approach does not prescribe the specific types of projects that must be

included in the Action Plan; Consortium partners have been encouraged to pursue the most cost-effective

options to achieve the agreed-upon goals for nitrogen management. The TBEP will review and revise

nitrogen management goals every five years, or more often if significant new information becomes

available.

SUMMARY

The Tampa Bay management community has agreed that protection and restoration of Tampa Bay living

resources is of primary importance. Through the TBEP process (initiated in 1991), partners have adopted

nitrogen loading targets for Tampa Bay based on the water quality requirements of Thalassia testudinum

and other native seagrass species. A long-term goal has been adopted to achieve 15,400 ha of seagrass in

Tampa Bay, or 95% of that observed in 1950. To reach the long-term seagrass restoration goal, a 7%

increase in nitrogen loading associated with a projected 20% increase in the watershed’s human

population over the next 20 years must be offset. Government and agency partners in the Tampa Bay

Estuary Program and private industries and interests participating in the Nitrogen Management

Consortium have identified and implemented specific nitrogen load reduction projects to ensure that

water quality conditions necessary to meet long-term living resource restoration goals for Tampa Bay are

achieved.

For more information:

National Research Council, National Academy of Sciences. 2000. Clean Coastal Waters:

Understanding and Reducing the Effects of Nutrient Pollution. National Academy Press, Washington,

D.C. 405 pages.

Janicki, A. and D. Wade. 1996. Estimating critical nitrogen loads for the Tampa Bay estuary: An

empirically based approach to setting management targets. Technical Publication 06-96 of the Tampa

Bay National Estuary Program. Coastal Environmental, Inc., St. Petersburg, Florida.
CS-26 Nutrient Criteria—Estuarine and Coastal Water

Johansson, J.O.R. and H. Greening. 2000. Seagrass restoration in Tampa Bay: A resource-based

approach to estuarine management. IN: Subtropical and Tropical Seagrass Management Ecology,

Bortone, S. (Ed.). CRC Press, Boca Raton, FL. 

Tampa Bay Estuary Program. 1998. Partnership for Progress: The Tampa Bay Nitrogen Management

Consortium Action Plan 1995-1999. Tampa Bay Estuary Program, St. Petersburg, Florida.
Nutrient Criteria—Estuarine and Coastal Water CS-27

CASE STUDY

RESTORING CHESAPEAKE BAY WATER QUALITY

Contact: Richard Batiuk, 410/267-5731; batiuk.richard@epa.gov

Original Nutrient Reduction Goal

In 1987, the Chesapeake Bay Program partners set a 40 percent reduction goal for nitrogen and

phosphorus to improve low oxygen conditions in the deep trench of the mainstem bay. The goal was

later defined to apply only to “controllable” sources, and only from the States—Maryland, Virginia,

Pennsylvania—and the District of Columbia are also listed as impaired tidal waters.

All listed impaired waters are scheduled to have a Total Maximum Daily Load or TMDL developed. A

TMDL defines the pollutant load that a waterbody can assimilate without causing violations of water

quality standards and allocates the loading to contributing point sources and nonpoint source categories. 

Once a TMDL is established by a State and approved by EPA through regulatory action, it is

implemented through regulatory and nonregulatory programs. A regulatory TMDL covering the entire

64,000 square mile bay watershed will be put in place by 2011 if bay water quality is not restored. 

Keeping A Cooperative Approach to Bay Restoration

To avoid potential negative impacts that a regulatory TMDL process might have on the successful,

cooperative efforts being used by the States’ tributary strategy programs, the Chesapeake 2000

Agreement lays out a series of commitments directed toward seeking a cooperative solution to restoring

bay water quality by 2010.

The bay watershed partners will define the water quality conditions necessary to support bay living

resources—fish, crabs, oyster, and bay grasses by 2001. These required conditions will be defined

through a series of Chesapeake Bay water quality criteria for dissolved oxygen, water clarity, and

chlorophyll a currently under development.

Important distinct bay and tributary tidal water habitats are being identified and characterized as

designated uses, where the above bay criteria will be applied to fully protect the aquatic living resources.

The States with bay tidal waters—Maryland, Virginia, Delaware, and the District of Columbia—have all

committed to adopting these bay criteria and tidal water designated uses into their individual State water

quality standards by 2003.

Critical to supporting the States’ adoption of the bay criteria and refined tidal waters designated uses will

be a baywide Use Attainability Analysis (UAA). 

Loading caps on nutrients and sediments needed to meet the bay water quality criteria will be allocated to

major tributary basins and individual States within those basins by December 2001.

Tributary strategies, detailed implementation plans to reach the allocated loading caps will be developed

in cooperation with local watershed stakeholders.

A reevaluation planned for 2005 will provide an opportunity for any necessary mid-course corrections on

the road to restoring bay water quality by 2010.
CS-28 Nutrient Criteria—Estuarine and Coastal Water

Bay Criteria: Defining Restored Bay Water Quality

The Chesapeake 2000 Agreement committed the signatories to the following: “by 2001, define the water

quality conditions necessary to protect aquatic living resources.” These water quality conditions are

being defined through the development of Chesapeake Bay specific water quality criteria for dissolved

oxygen, water clarity, and chlorophyll a. Collectively, these three water quality parameters provide the

best and most direct measures of the impacts of too much nutrient and sediment pollution on the bay’s

aquatic living resources—fish, crabs, oysters, and underwater bay grasses. 

Bay Criteria

Dissolved Oxygen

Fish and other aquatic life require levels of dissolved oxygen to survive. Seasonal algae blooms deplete

dissolved oxygen, potentially rendering deep waters of the bay uninhabitable to certain species, such as

the endangered Atlantic Sturgeon during certain times of the year. Bay dissolved oxygen levels should

be those required by the aquatic communities inhabiting different parts of the bay during different times

of the year, fully reflective of natural conditions.

Chlorophyll a

Measurements of chlorophyll indicate levels of phytoplankton or algal biomass in the water column. Bay

chlorophyll levels should be moderate: not so high as to cause harmful algal blooms that lead to poor

quality food, shading of light in shallow water habitats, and low dissolved oxygen conditions when the

algae die off and sink to the bottom.

Water Clarity

Underwater grasses collectively are an essential component of the bay’s living resources habitat. 

Decreased water clarity inhibits the growth of underwater bay grasses. Water clarity is adversely affected

by increased sediment loads and algal biomass spurred by excess nutrient inputs to the bay. Bay water

quality conditions should generally provide high water clarity—sunlight penetration—to support

restoration of underwater grasses throughout the bay’s extensive shallow water habitats.

Chesapeake Bay Dissolved Oxygen Criteria

Chesapeake Bay Dissolved Oxygen Dynamics

The Chesapeake Bay has a built-in, natural tendency toward reduced dissolved oxygen conditions,

particularly within its deeper waters because of the physical morphology and estuarine circulation. Its

highly productive, shallow waters, coupled with its tendency to retain, recycle, and regenerate the

nutrients delivered from the atmosphere and surrounding watershed set the stage for a nutrient-rich

environment. The mainstem Chesapeake Bay and its major tidal rivers with deep channels coming off

shallower, broad shoal waters, and the significant influx of freshwater flows result in stratification of the

water column, essentially locking off deeper bottom waters from mixing with higher oxygenated surface

waters. Combined together, the retention/efficient recycling of nutrients and water column stratification

lead to severe reductions in dissolved oxygen concentrations during the warmer months of the year,

generally May to September.

Nearshore, shallow waters in the Chesapeake Bay also periodically experience episodes of low to no

dissolved oxygen conditions, in part, resulting from intrusions of bottom water forced onto the shallow

flanks by sustained winds (Carter et al.1978, Tyler 1984, Seilger et al.1985, Malone et al.1986). Diel

cycles of low dissolved oxygen conditions often occur in nonstratified shallow waters where nighttime

water column respiration temporarily depletes dissolved oxygen levels (D’Avanzo and Kremer 1994).

The timing and spatial and volumetric extent of hypoxic and anoxic waters vary from year to year,

largely driven by local weather patterns, timing and magnitude of freshwater river flow and concurrent
1

 These target species were from a larger list of commercially, recreationally , and ecologically important species

reported in Habitat Requirements for Chesapeake Bay Living Resources-Second Edition (Funderburk et al. 1991).

Nutrient Criteria—Estuarine and Coastal Water CS-29

delivery of nutrients and sediments into tidal waters, and the corresponding springtime phytoplankton

bloom (Officer et al.1984, Seliger et al.1985). In Chesapeake Bay mainstem, the onset of low to no

dissolved oxygen conditions can be as early as April and persist through September, until fall turnover of

the water column. The deeper waters of major tidal tributaries can exhibit hypoxic and anoxic

conditions, with the nature, extent, and magnitude of low dissolved oxygen and the causative factors

varying from river to river.

The scientific underpinnings of these Chesapeake Bay specific criteria have been in the works for

decades. Seasonal low dissolved oxygen conditions in the Chesapeake Bay were first documented in the

1930s (Newcombe and Horne 1938). Basic understanding of dissolved oxygen dynamics, critical to

derivation of criteria reflective of ecosystem process, began with the research cruises of the Chesapeake

Bay Institute from the 1950s through the late 1970s. A 5-year multidisciplinary research program

starting in the late 1980s, funded by the Maryland and Virginia Sea Grant Program, yielded significant

advances in understanding of all facets of oxygen dynamics, effects, and ecosystem implications (Smith

et al. 1991). These investigations laid the groundwork for more management-focused applications of the

science.

Chesapeake Bay Dissolved Oxygen Restoration Goal

Published in 1992, the Chesapeake Bay dissolved oxygen restoration goal was developed in response to

the Chesapeake Executive Council’s commitment “to develop and adopt guidelines for the protection of

water quality and habitat conditions necessary to support the living resources found in the Chesapeake

Bay system and to use these guidelines.” The dissolved oxygen restoration goal consisted of a narrative

statement supported by specific target dissolved oxygen concentrations applied over specified averaging

periods and locations. Dissolved oxygen effects information was compiled for 14 identified target

species1 of fish, molluscs, and crustaceans as well as for other supporting benthic and planktonic species

within the bay food web. The target concentrations and their specified temporal averaging and spatial

application were determined from analysis of dissolved oxygen levels that would provide the levels of

protection described within the narrative restoration goal. Best professional judgment was used in areas

where there were gaps in the information base on dissolved oxygen effects available a decade ago.

The original dissolved oxygen restoration goal and its supporting framework made three breakthroughs at

that time of significance to supporting derivation and management application of the Chesapeake Bay

specific dissolved oxygen criteria within this document. The dissolved oxygen target concentrations

varied with vertical depth through the water column as well as horizontally across the expanse of the bay

and its tidal tributaries, directly reflecting variations in required levels of protection for different living

resource habitats. The averaging periods for each target concentration were tailored to specific habitats,

recognizing that short-term exposures to concentrations below the target concentrations were allowable

and still protective of living resources. The dissolved oxygen goal document contained a methodology

through which water quality monitoring data and model-simulated outputs collected over varying

frequencies could be directly assessed in terms of the percentage of time that areas of bottom habitat or

volumes of water column habitat were predicted to meet or exceed the applicable target dissolved oxygen

concentrations.

Regionalizing the EPA Virginian Province Saltwater Dissolved Oxygen Criteria

With the publication of the EPA Ambient Water Quality Criteria for Dissolved Oxygen (Saltwater): Cape

Cod to Cape Hatteras came a decade’s worth of systematically developed dissolved oxygen effect data

along with synthesis and close evaluation of several decades of effects data published in the scientific
CS-30 Nutrient Criteria—Estuarine and Coastal Water

peer reviewed literature (Thursby et al.2000). The approach to derive these dissolved oxygen criteria

combined features of the traditional water quality criteria with a new biological framework. A

mathematical model was used to integrate time (replacing the concept of an averaging period) and

establish protection limits for different life stages (i.e., larvae verus juveniles and adults). Where

practical, data were selected and analyzed in manners consistent with the Guidelines for Deriving

Numerical National Water Quality Criteria for the Protection of Aquatic Organisms and Their Uses

(hereafter referred to as the EPA Guidelines) (Stephan et al.1985).

The EPA Virginian Province dissolved oxygen saltwater criteria document addressed three areas of

protection: (1) juvenile and adult survival, (2) growth effects, and (3) larval recruitment effects. In doing

so, the criteria document segregated effects on juveniles and adults from those on larvae. The survival

data on the sensitivity of the juveniles and adults are handled in a traditional EPA guidelines manner. To

address cumulative effects of low dissolved oxygen on larval recruitment to the juvenile life stage (i.e.,

larval survival as a function of time) a new biological approach was taken. These criteria were derived

using a mathematical model that evaluates the effect of dissolved oxygen conditions on larvae by

tracking the intensity and duration of low dissolved oxygen effects across the larval recruitment season. 

Protection of larvae of all species is provided by using low dissolved oxygen effects data on larval stages

of nine sensitive estuarine/coastal organisms.

The Virginian Province saltwater dissolved oxygen criteria document and its underlying effects database

and methodologies were structured to support regional specific derivation of dissolved oxygen criteria

tailored to the species, habitats, and nature of dissolved oxygen exposure regimes of different estuarine,

coastal, and marine waters. The segregation by life stages allows the criteria to be factored into the

refined tidal water designated uses, which themselves, in part, reflect use of different habitats by

different life stages. This segregation by life stage is a significant difference from traditional aquatic life

criteria.

However, the Virginian Province saltwater criteria were not explicitly set up to address natural vertical

variations in dissolved oxygen concentration. If Chesapeake Bay specific criteria were derived through a

strict application of the EPA saltwater criteria methodology, there would not be the flexibility needed to

tailor each set of criteria to the refined tidal water designated uses. The resultant bay criteria would be

driven solely by larval effects data irrespective of depth and season.

The Chesapeake Bay specific criteria were derived through the regional application of the Virginian

Province effects database and application of traditional toxicological and new biological-based criteria

derivation methodologies. Chesapeake Bay specific science was factored directly into each step of the

criteria derivation process. The extensive Virginian Province dissolved oxygen effects database was first

focused down on only Chesapeake Bay species and then supplemented with additional Chesapeake Bay

species effects data from the scientific literature. The Virginian Province larval recruitment model was

modified to better reflect Chesapeake Bay conditions, with its application broadened to include

additional Chesapeake Bay species. Finally, specific steps were taken to factor the requirement to

provide protection of species listed as threatened/endangered in Chesapeake Bay into the bay-specific

criteria.

Current State water quality standards generally require 5 mg/L of dissolved oxygen throughout all of the

bay’s waters—from the deep trench near the bay’s mouth to the shallows at the head of the bay. Even

though the 5 mg/L standard is baywide, bay region scientists believe natural conditions dictate that in

some sections of the bay, such as the deep channel, bay waters cannot achieve the current 5 mg/L

standard during the warmer months of the year. Additionally, scientists believe other areas, such as

prime migratory fish spawning areas, require higher levels of dissolved oxygen to sustain life during the 
Nutrient Criteria—Estuarine and Coastal Water CS-31

38

CBP 8/9/01

CHESAPEAKE BAY PROGRAM

Dissolved Oxygen 

Criteria

Section 4: How Will We Define Restored Water Quality?

Minimum Amount of Oxygen (mg/L) 

Needed to Survive by Species

Hard Clams: 5

Striped Bass: 5-6

Worms: 1

Migratory Spawning & 

Nursery Areas

Shallow and Open Water 

Areas

Deep Water

Deep Channel

6

5

3

2

1

4

0

Crabs: 3

Spot: 2

White Perch: 5

American Shad: 5

Yellow Perch: 5

Alewife: 3.6

Bay Anchovy: 3

Figure 1. Dissolved Oxygen Criteria, Chesapeake Bay.

late winter to early summer timeframe. The amount of oxygen needed in the bay tidal waters depends on

specific needs of the aquatic living resources and where they live and during which time of the year they

live there.

The Chesapeake Bay dissolved oxygen criteria vary significantly across the five proposed tidal water

designated uses to fully reflect the wide array of species living in these different bay habitats (Figure 1). 

These working draft dissolved oxygen criteria were developed by the Chesapeake Bay Dissolved Oxygen

Criteria Team, a bay region team composed of scientists, State and federal managers, and technical

stakeholders (Table 1). A draft document describing the Chesapeake Bay Dissolved Oxygen Criteria in

greater detail is available for review and comment at www.chesapeakebay.net. There is a year-long

process and schedule, including three public reviews, leading to publication of these Chesapeake Bay

specific water quality criteria by EPA by June 2002.

Chesapeake Bay Chlorophyll a Criteria

Chlorophyll a is used to measure the abundance and variety of microscopic plants or algae that form the

base of the food chain in the bay. Excessive nutrients can stimulate nuisance algae blooms, resulting in

reduced water clarity, reduced amount of good quality food and depleted oxygen levels in deeper water. 

By its very nature, chlorophyll a is both an integrated biological measure of production of the primary

food source of the entire bay food web as well as a critical indicator of water quality through its direct

role in reducing light penetration and fueling bacterial processes leading to low dissolved oxygen levels. 

As stated upfront by Harding and Perry (1997), “chlorophyll a is a useful expression of phytoplankton

biomass and is arguably the single most responsive indicator of N [nitrogen] and P [phosphorus]

enrichment in this system [Chesapeake Bay].” Determining the levels of chlorophyll a which are fully

protective of the refined designated uses of the vast tidal waters that compose the Chesapeake Bay and

tributaries must factor in all the different roles chlorophyll a plays in defining a restored, more balanced

Chesapeake Bay ecosystem.
CS-32 Nutrient Criteria—Estuarine and Coastal Water

Table 1. Working Draft Chesapeake Bay Dissolved Oxygen Criteria (July 3, 2001)

Designated Use Criteria Concentration/Duration Temporal Application

Migratory spawning and

nursery

7 day mean of 6 mg/La February 15th - June 10th

Instantaneous minimum of 5 mg/L

30 day mean of 5 mg/L

June 11th - February 14th

7 day mean of 4 mg/L

Instantaneous minimum of 3.5 mg/L

Shallow/open water 

30 day mean of 5 mg/L

All year round

7 day mean of 4 mg/L

Instantaneous minimum of 3.5 mg/L

Deeper water 30 day mean of 3 mg/L April through September

Instantaneous minimum of 1.7 mg/L

30 day mean of 5 mg/L

October through March

7 day mean of 4 mg/L

Instantaneous minimum of 3.5 mg/L

Deep channel

Instantaneous minimum of 1 mg/L April through September

30 day mean of 5 mg/L

October through March

7 day mean of 4 mg/L

Instantaneous minimum of 3.5 mg/L

a

 Applied to tidal fresh waters with long term averaged salinities less than 0.5 parts per thousand.

The derivation of the Chesapeake Bay chlorophyll a criteria were based on the convergence of several

independent lines of evidence—historical observed concentrations, literature values related to trophic

status, direct contributions to light attenuation, and contribution to dissolved oxygen

conditions—collaborating chlorophyll a concentrations derived as a result of characterizing set of

phytoplankton reference communities.

Phytoplankton Reference Community/Food Quality Connection

Estimates of phytoplankton taxon biomasses were derived from the Maryland and Virginia Chesapeake

Bay Monitoring Program phytoplankton count data (1984-1999) and along with other phytoplankton

indicators—chlorophyll a, pheophytin, and primary productivity—were used to investigate differences in

biomass, taxonomic composition, and food value for the range of water quality conditions currently

experienced in the Chesapeake Bay. The biological data were sorted into categories based on seasonand salinity-specific concentrations/levels of three parameters in the associated water quality data:

dissolved inorganic nitrogen, ortho-phosphate and Secchi depth. Relatively small secchi depths and

excess dissolved inorganic nitrogen and excess ortho-phosphate characterized the Poor water quality

categories. Relatively high light levels and algal growth-limiting concentrations of dissolved inorganic

nitrogen and ortho-phosphate characterized the good water quality categories. Mixed water quality

conditions (i.e., one or two water quality parameters qualified as Better but the other(s) did not) and

extreme subsets of the Poor and Better categories (i.e. Worst and Best) were also investigated. 

Qualitative and quantitative measures of the phytoplankton community composition and biomass
Nutrient Criteria—Estuarine and Coastal Water CS-33

distributions were then evaluated relative to these water quality classifications and implications for food

quality and quantity for filter feeding fish and shellfish.

Historically Observed Concentrations

Several recent in-depth reviews and evaluations of historically observed (1950s to early 1980s) and

current (1984-1998) chlorophyll a concentrations provide a strong basis for collaborating the Chesapeake

Bay specific chlorophyll a criteria (Harding 1994, Harding and Perry 1997, Olson and Lacouture, in

review). Using information from five decades of water quality data provides insights into both

chlorophyll a concentrations that are attainable under a range of otherwise natural conditions

(meteorological, river flow, tidal flushing) as well as concentrations reflective of a healthier bay

ecosystem. 

Literature Values Related to Trophic Status

Throughout the scientific literature, there are several defining papers which through synthesis of a wide

array of data from many different aquatic systems center down on ranges of conditions reflective of

different trophic states of water bodies (e.g., Wetzel 1985, Ryding and Rast 1989, Smith et al.1992). 

Chlorophyll a is a principal parameter quantified within these literature reviews. The strength of this

collaborative line of evidence is that information is drawn from diversity of systems across the spectrum

of healthy to clearly eutrophied water bodies. This approach provides insights into common

characteristics associated with trophic status that can not be drawn through the study of a single, although

large, water body like Chesapeake Bay.

Direct Contributions to Light Attenuation

Over the past four decades, the Chesapeake Bay ecosystem has had an extensive, widely distributed

underwater grass community undergo severe declines followed by a decade and a half slow but steady

recovery. The bay management and scientific communities have invested significant resources in the

investigation of this grand natural experiment, learning much about the causes of the decline and

potential solutions for continued, yet accelerated restoration. Two comprehensive technical synthesizes

of this wealth of scientific knowledge and insights have been published which provide direct quantitative

insights into the role of chlorophyll a in the recovery of underwater bay grasses (Batiuk et al.1992,

2000). This collaborative line of evidence draws on the chlorophyll a connection to reductions in light

penetration through the water column.

Contribution to Dissolved Oxygen Conditions

It is well known and documented that algae uneaten by higher trophic levels—zooplankton, oysters and

fish of all kinds—becomes the fuel, through its breakdown by bacteria, for reducing dissolved oxygen

levels. Through an analysis of Chesapeake Bay water quality model simulated outputs from scenarios

which simulated dissolved oxygen conditions which met the dissolved oxygen criteria, the model

simulated chlorophyll a concentrations of desired dissolved oxygen conditions were quantified.

Appropriate chlorophyll a levels vary, depending on the salinity of the water. The proposed criteria for

chlorophyll a are split out from tidal freshwater all the way to very salty—polyhaline—waters. Season

of the year is also important, with spring and summer being the most important times of year that high

chlorophyll a levels can impact living resources in the bay. 

These working draft chlorophyll a criteria were developed by the Chesapeake Bay Chlorophyll and

Nutrient Criteria Team, a bay region team composed of scientists, State and federal managers, and

technical stakeholders (Table 2). A draft document describing the Chesapeake Bay Chlorophyll a

Criteria in greater detail is available for review and comment at www.chesapeakebay.net. There is a 
CS-34 Nutrient Criteria—Estuarine and Coastal Water

Table 2. Working Draft Chesapeake Bay Chlorophyll a Criteria (July 3, 2001)

Salinity Regime

Chesapeake Bay Chlorophyll Criteria (ug/L)

Spring (March-May) Summer (July-September)

Median Maximum Median Maximum

Tidal Fresh 8 12 9 16

Oligohaline 10 23 6 23

Mesohaline 6 27 7 16

Polyhaline 3 7 4 9

year-long process and schedule, including three public reviews, leading to publication of these

Chesapeake Bay specific water quality criteria by EPA by June 2002.

Connection to Underwater Bay Grasses

The loss of submerged aquatic vegetation, or SAV, from shallow waters of Chesapeake Bay, which was

first noted in the early 1960s, is a widespread, well-documented problem. Although other factors, such

as climatic events and herbicide toxicity, may have contributed to the decline of SAV in the bay, the

primary causes are eutrophication and associated reductions in light availability. The loss of SAV beds

are of particular concern because these plants create rich animal habitats that support the growth of

diverse fish and invertebrate populations. Similar declines in SAV have been occurring worldwide with

increasing frequency during the last several decades. Many of these declines have been attributed to

excessive nutrient enrichment and decreases in light availability. 

Chesapeake Bay Water Clarity Criteria

One of the major features contributing to the high productivity of Chesapeake Bay has been the historical

abundance of SAV. There are over 20 freshwater and marine species of rooted, submerged flowering

plants in bay tidal waters. These underwater grasses provide food for waterfowl and are critical habitat

for shellfish and finfish. SAV also affect nutrient cycling, sediment stability, and water turbidity.

The health and survival of these plant communities in Chesapeake Bay and its tidal tributaries depend on

suitable environmental conditions that define the quality of SAV habitat. Key to the restoration of these

critical habitats and food sources is the return of levels of light penetration in shallow waters necessary to

support the survival, growth, and repropagation of diverse, healthy underwater bay grass communities.

Bay Water Clarity Derivation Approach

Through the combined efforts of the bay’s scientific and resource management communities, two

internationally recognized technical syntheses of information supporting quantitative habitat

requirements for Chesapeake Bay SAV have been published in the past decade (Batiuk et al.1992, Batiuk

et al. 2000). Key findings, the underlying light requirements, and management-oriented diagnostic tools

and restoration targets have been reported in the peer reviewed scientific literature (Dennison et al. 1993,

Kemp et al. in review; Gallegos 2001, Koch 2001, Bergstrom in preparation, Carter and Rybicki in

preparation, Karrh in preparation, Kemp et al. in preparation). These two technical syntheses of

worldwide literature, bay-specific research and field studies, and recent model simulation and data

evaluation provide the scientific foundation for the Chesapeake Bay water clarity criteria described here. 

Readers are encouraged to consult these two syntheses and the resultant scientific literature papers for

more in-depth technical details and documentation.
Nutrient Criteria—Estuarine and Coastal Water CS-35

36

CBP 8/9/01

CHESAPEAKE BAY PROGRAM

What’s Blocking the Light?

Good Water Clarity Poor Water Clarity

Algae in the water

Sediment and other 

particles in the water

Algae on the leaves

Percent of sunlight 

reaching leaves:

•9% in low salinity 

waters

•15% in high salinity 

waters

+

+

equals

Very low percentage 

of sunlight reaching 

leaves – Bay grasses 

grow poorly or die.

Section 4: How Will We Define Restored Water Quality?

Figure 2. Bay Water Clarity.

The Chesapeake Bay specific water clarity criteria derivation follows four successive stages: first,

determination of water-column-based light requirements for SAV survival and growth, then

quantification of the factors contributing to water column light attenuation. The contributions from

epiphytes to light attenuation at the leaf surface are then factored into methods for estimating total light

attenuation. Finally, a set of minimal light requirements are determined as the actual criteria values.

The draft bay criteria propose that water clarity criteria should apply to areas of the bay that are up to 2

meters deep (approximately 6 feet). Areas where SAV never occurred historically, or where natural

factors prevent its growth (e.g., strong currents, rocky bottoms) would be excluded. The water clarity

criteria reflect the different light requirements for underwater plant communities that inhabit low salinity

versus higher salinity shallow water habitats throughout the bay (Figure 2).

These working draft water clarity criteria were developed by the Chesapeake Bay Water Clarity Criteria

Team, a bay region team composed of scientists, State and federal managers, and technical stakeholders

(Table 3). A draft document describing the Chesapeake Bay Water Clarity Criteria in greater detail is

available for review and comment at www.chesapeakebay.net. There is a year-long process and

schedule, including three public reviews, leading to publication of these Chesapeake Bay specific water

quality criteria by EPA by June 2002.

These Chesapeake Bay criteria will be applied to a series of designated uses which, in turn, reflect key

habitats throughout the bay and its tidal tributaries.

Excessive Nutrient and Sediment Loads

The causes of these water quality impairments—excessive loadings of nitrogen, phosphorus, and

sediment—will be addressed through commitments to determine reductions in loadings needed to

achieve the bay criteria. These required loading reductions will be established as caps on loadings 
CS-36 Nutrient Criteria—Estuarine and Coastal Water

Table 3. Working Draft Chesapeake Bay Water Clarity Criteria (July 3, 2001)

Habitat Category Criteria Concentration

(percent ambient light)

Temporal Application

Tidal fresh shallow water 9 % April - October

Oligohaline shallow water 9 % April - October

Mesohaline shallow water 15 % April - October

Polyhaline shallow water 15% March-May, Sept-Nov.

allocated to each tributary basin within the Chesapeake Bay watershed. This approach is consistent with

EPA’s regional establishment of ambient concentration-based nutrient criteria, but places more emphasis

on the water quality parameters with a direct impact on aquatic living resources. Through this approach

nutrients and sediments are addressed directly through caps on loading determined through application of

the linked bay airshed-watershed-tidal water quality models and analysis of Chesapeake Bay Monitoring

Program data in place of the development of ambient nutrient and sediment criteria.

Relationship with National Efforts to Develop Nutrient Criteria

At the same time, a parallel effort is currently underway by EPA to develop ecoregion specific numerical

nutrient criteria across the country to meet the objectives of the Clean Water Action Plan. A nutrient

criteria team has been established by EPA-Region III to implement the National Nutrient Strategy issued

by EPA last year for the mid-Atlantic region. The EPA Region III team is focusing its nutrient criteria

development efforts on the free flowing stream, rivers, lakes, and wetlands within the mid-Atlantic

States, not the Chesapeake Bay tidal waters. Whereas the bay criteria are focused on dissolved oxygen,

water clarity and chlorophyll a, the EPA Region III team is developing ambient concentration criteria for

total nitrogen, total phosphorus, chlorophyll, and turbidity.

The advanced scientific understanding of water quality impacts on aquatic bay living resources combined

with the state of the art linked bay airshed-watershed -water quality models enabled the bay watershed

partners to develop criteria for water quality measures directly influencing aquatic resources. The cause

of reduced water quality conditions—too much nitrogen, phosphorus, and sediment—will be addressed

through the establishment of loading caps. The bay models enable the partners to effectively translate the

desired dissolved oxygen, water clarity, and chlorophyll a conditions back into reduced loadings of

nutrients and sediments from the surrounding watershed and airshed. Bay science has shown that it is the

delivered loads of nutrients and sediment, not just the ambient concentrations, that have had an impact on

oxygen, light, and algae levels in the bay tidal waters.

Bay Tidal Water Designated Uses

Because conditions throughout the Chesapeake Bay tidal water habitats differ based on depth, salinity

and season, a uniform baywide water quality standard does not take into account the varying needs of

different plants and animals. As a result, current State water quality standards, which differ between the

four jurisdictions with tidal waters, need to be revised and expanded to account for the natural variability

in conditions found throughout the bay. Each of the three bay criteria will differ from one region of the

bay and its tidal tributaries to another, as determined by the plants and animals residing in that area. 

Once the bay criteria and the tidal water designated uses are adopted as State water quality standards,

these tailored set of standards will apply to similar habitats across all jurisdictions. 

An area’s designated use refers to a waterbody’s function—such as fishable or swimable—and takes into

account the use of the water body for public water supply, the protection of fish, shellfish, and wildlife,

as well as its recreational, agricultural, industrial and navigational purposes. The existing Maryland,
Nutrient Criteria—Estuarine and Coastal Water CS-37

Refined Designated Uses for

Chesapeake Bay and Tidal Tributary Waters

A. Cross Section of Chesapeake Bay or Tidal Tributary

B. Oblique View of the “Chesapeake Bay” and its Tidal Tributaries

Shallow Water

Open Water

Deep Water

Deep Channel

Open Water

Shallow Water Habitat

Habitat

Deep Water

Deep Channel

Migratory Finfish

Spawning and

Nursery Habitat

Figure 3. Refined Designated Uses for Chesapeake Bay Tidal Tributary Waters.

Virginia, Delaware, and District of Columbia designated uses for the bay’s tidal waters do not fully

reflect the wide variety of different habitats found throughout the bay and its tidal tributaries. Where two

jurisdictions boundaries join, each State has different designated uses for the same waterbodies.

Five refined Chesapeake Bay tidal water designated uses have been established to more fully reflect the

different aquatic living resource communities inhabiting a variety habitats and, therefore, the different

intended aquatic life uses of those tidal habitats (Figure 3).

The Migratory Spawning & Nursery designated use is the propagation and growth of balanced

indigenous populations of ecologically, recreationally, and commercially important anadromous, semianadromous, and tidal fresh resident fish species inhabiting spawning and nursery grounds. 

The Shallow Water designated use is the propagation and growth of balanced, indigenous populations of

ecologically, recreationally, and commercially important fish, shellfish and underwater grasses inhabiting

shallow waters habitats.

The Open Water designated use is the propagation and growth of balanced, indigenous populations of

ecologically, recreationally, and commercially important fish, and shellfish species that inhabit open

water habitats. 

The Deep Water designated use is the propagation and growth of balanced, indigenous populations of

ecologically, recreationally, and commercially important fish and shellfish species inhabiting deep water

habitats. 

The Deep Channel designated use is to provide a refuge for balanced, indigenous populations of

ecologically, recreationally, and commercially important fish species that depend on deep channel
CS-38 Nutrient Criteria—Estuarine and Coastal Water

Deep Channel U

Deep Water U

Open Water U U

Shallow Water U U U

Migratory U U

Spawning and 

Nursery

Dissolved Chlorophyll a Water Clarity

Oxygen

Table 4. Chesapeake Bay Criteria Needed for Protection

of the Proposed Tidal Waters Designated Uses

habitats for overwintering during colder months of the year and the propagation and growth of benthic

infaunal and epifaunal worms and clams that provide food for bottom feeding fish and crabs. 

These tidal water designated uses were developed by the Chesapeake Bay Water Quality Standards

Coordinators Team, a bay region team composed of water quality standards coordinators from all six

States, the District of Columbia, EPA Region 2, 3, and headquarters offices. Table 4 shows how refined

tidal water designated uses relate to the bay criteria. 

The watershed partners are evaluating the refined tidal water designated uses and the applicable bay

criteria through a baywide use attainability analysis. The final tidal water designated uses will be

adopted by Maryland, Virginia, Delaware, and the District of Columbia, along with the applicable bay

water quality criteria into their State water quality standards by 2003. These refined designated uses will

add more specifics to the existing State designated uses and apply consistently across jurisdictions for

similar habitats.

Baywide Use Attainability Analysis

The Chesapeake 2000 Agreement commits the States with bay tidal waters—Maryland, Virginia, and

Delaware—and the District of Columbia to adopt into their State water quality standards as consistent set

of bay criteria and designated uses across bay tidal habitats. Whenever there is a proposed change in

water quality standards, such as that being undertaken for Chesapeake Bay waters, it is necessary to

assess attainment of the designated uses and underlying criteria. Such an assessment is called a Use

Attainability Analysis or UAA.

A UAA is used by States to justify changes to their water quality standards by assessing the physical,

chemical, biological, economic, or other factors affecting attainment of the designated use. The UAA

describes the scientific attributes of the waterbody, both natural conditions and conditions brought about

by human contribution. If the attributes of the waterbody make attaining the use impossible, or if there

are economic reasons why the use cannot be attained, the UAA is used to clearly document these reasons. 

Finally, the UAA describes how the proposed standards will protect existing uses. All six bay watershed

States—New York, Pennsylvania, Maryland, Virginia, West Virginia and Delaware—along with the

District of Columbia and EPA are working together with bay watershed partners to carry out such an
Nutrient Criteria—Estuarine and Coastal Water CS-39

assessment. A use attainment assessment on a scale as large as the 64,000 square mile Chesapeake Bay

watershed has never been carried out.

Adopting Bay Criteria as State Water Quality Standards

Water quality standards combine water quality criteria and designated uses to produce a target numeric

value assigned to a waterbody that, if achieved, will maintain healthy water quality. Through the

Chesapeake 2000 Agreement, Maryland, Virginia, and the District of Columbia are committed to

adopting the new bay criteria–dissolved oxygen, water clarity, and chlorophyll a—along with the refined

tidal water designated uses as State water quality standards. Delaware, which shares bay tidal waters

with Maryland in its portion of the Nanticoke River watershed, has made the same commitment through

the six-State memorandum of agreement. Together, the States and the District must achieve these new

bay-specific water quality standards needed to support restored estuarine ecosystem if the Chesapeake

Bay is to be removed from the list of impaired waters. 

Why New State Standards

Existing State water quality standards are applied broadly across each State’s tidal waters, without

recognition of the variety of habitats. Each State has different water quality standards applied to the

same tidal waters, whereas the aquatic living resources in bay habitats, which do not recognize these

jurisdictional boundaries, may have the same water quality needs. Currently, dissolved oxygen is the

only numerical water quality standard adopted by all three States and the District of Columbia that

addresses nutrient- and sediment-related water quality pollution problems.

So compliance with existing State water quality standards will not fully protect the living resources in the

bay waters. In some critical habitats of the bay, specifically migratory fish spawning and nursery areas,

existing State water quality standards will not fully protect more sensitive life stages. In other cases,

reaching existing standards is not possible owing to natural conditions found in deeper bay waters during

the warmer months of the year. Existing State water quality standards do not include measures to protect

underwater bay grasses or fully support good quality fish food.

Setting and Allocating New Cap Loads

The Chesapeake 2000 Agreement commits the signatories to determining the nutrient and sediment load

reductions necessary to achieve the water quality conditions that protect aquatic living resources. Those

load reductions will then be assigned or allocated to each major tributary basin in the form of cap loads. 

Cap loads are the maximum amounts of pollutants allowed to flow into a waterbody and still ensure

achievement of State water quality standards. In this case, the water quality standards will be the new

bay criteria and refined tidal water designated uses (currently in draft) to be adopted by the States of

Maryland, Virginia, and Delaware, and the District of Columbia into their standards by 2003. 

Available Tools and Information

The Chesapeake Bay watershed partners will use the Chesapeake Bay Airshed, Watershed and Estuary

Models, the USGS SPARROW model, along with Chesapeake Bay Monitoring Program data, to help

determine these cap loads for nitrogen, phosphorus, and sediment. These models are mathematical

representations that simulate the real world, interpreting various levels of actions (management

scenarios) to reduce different amounts of pollutant loads. These scenarios are run through the models to

determine how to achieve baywide attainment of the bay water quality criteria for dissolved oxygen,

water clarity, and chlorophyll a as applied to the tidal water designated uses. 

Cap Setting and Allocation

These models and other available information will be used to allocate loading caps to the nine major

tributary basins–Susquehanna, Upper Western Shore, Patuxent, Potomac, Rappahannock, York, James,
CS-40 Nutrient Criteria—Estuarine and Coastal Water

Figure 4. The Nine Major Basins.

Upper Eastern Shore, and Virginia Eastern Shore (Figure 4). Each State and the District will bear a

proportional burden for achieving and maintaining the cap based on their existing pollutant loadings,

progress to date, effectiveness and cost efficiency considerations, and their pollutant loading effects on

different tributaries.

For multijurisdictional waters like the Susquehanna, Potomac, and Eastern Shore basins, the linked

watershed and bay water quality models will be used to further allocate cap load responsibilities to each

State.

Working with their local stakeholders, individual States to further subdivide their major tributary basin

load cap allocations into the 37 State-defined tributary strategy sub-basins. 

A comprehensive 2-year schedule has been set up to coordinate the efforts of the six watershed States,

the District of Columbia, and the many other involved bay watershed partners. The schedule also will

ensure direct and continued involvement of local stakeholders and the general public during the entire

cap load setting and allocation process. The Chesapeake Bay Water Quality Steering Committee,

composed of senior managers from the seven watershed jurisdictions, EPA regional and headquarters

offices, Chesapeake Bay Commission, river basin commissions, and involved stakeholders, has the

overall responsibility overseeing and reaching agreement on the cap load allocations. Many of the

subcommittees and workgroups within the Chesapeake Bay Program committee structure will be carrying

out the technical, modeling, data interpretation, economic analysis, policy evaluation, and

communication work in support of setting and allocating the nutrient and sediment cap loads.
Nutrient Criteria—Estuarine and Coastal Water CS-41

Tributary Strategies: Local Watershed Implementation 

The Chesapeake 2000 Agreement commits the bay watershed partners to “complete a public process to

develop and begin implementation of revised Tributary Strategies to achieve and maintain the assigned

loading goals.” Tributary strategies are detailed descriptions of planned local actions—riparian forest

buffer replanting, wastewater treatment upgrades, nutrient management on farms, stormwater treatment,

stream restoration, and many others—and a schedule for undertaking those actions necessary to reduce

nutrients and sediment loads from each tributary watershed to reach the assigned loading cap by 2010.

Development of tributary strategies has been a very driven public process with the direct involvement by

local governments, watershed associations, regional organizations, and a wide variety of other interested

local stakeholders. In creating the strategies, the States, and the District of Columbia work closely with

those groups and individuals within each respective watershed who will be directly involved in

implementation strategy. Together, they explore and evaluate a wide variety of point and nonpoint

source pollution control measures. They then draft a strategy using the most effective reduction options

to achieve the cap load allocated to their tributary strategy basin. 

The existing Tributary Strategies were designed to achieve the 1987 Bay Agreement goal of a 40 percent

reduction in nutrient loads from controllable sources from 1985 levels. Copies of these existing tributary

strategies are available on-line through the respective Maryland, Virginia, Pennsylvania, and the District

of Columbia tributary strategy web pages.

To restore the tidal water conditions necessary to sustain the bay’s fish, crabs, oysters and underwater

grasses will likely require greater reductions in nutrients in many areas then called for by the existing

tributary strategies. In addition, the water clarity conditions needed to restore underwater bay grasses

can not be achieved without significant reductions in sediments loads to the tidal waters. With New

York, Delaware and West Virginia joining as bay watershed partners through a six-State memorandum of

understanding, new tributary strategies will be developed for these States’ portions of the bay watershed

not previously addressed under the existing tributary strategies. 

The new and revised tributary strategies will now cover all sources of nutrient and sediment pollution,

including air sources, across the entire 64,000 Chesapeake Bay watershed. Tributary strategies will

address nutrient and sediment loading caps allocated to 37 sub-basins across the bay watershed by the

bay watershed partners. Loading reductions required through local stream and river segment regulatory

TMDLs will be directly integrated into each respective Tributary Strategy as part of the overall effort to

effectively “blend” the regulatory TMDL program and cooperative Chesapeake Bay Program.

Information on Local Watersheds

A wide array of information is available on local watersheds within the Chesapeake Bay basin, including

information directly relevant to the overall process for setting and allocating new cap loads through the

Chesapeake Watershed Profiles. Through this point and click information system linking the bay

watershed partners, one can access information on pollution sources, recent modeling results, status of

bay criteria attainment, long-term trends in water quality and living resources, draft cap allocations, and

more from the entire bay basin scale to local watersheds.
CS-42 Nutrient Criteria—Estuarine and Coastal Water

CASE STUDY

A PERSPECTIVE FROM WASHINGTON STATE

Jan Newton, Washington State Department of Ecology; Randy Shuman, King County Department of

Natural Resources; Greg Pelletier, Washington State Department of Ecology

The issue of nutrient control for marine receiving waters in Washington State has its origin in a landmark

case concerning freshwater eutrophication and lake restoration. In the early 1950's, Lake Washington, a

large lake (85 km2) situated near Seattle, Washington, was showing warning signs of ecological

deterioration. Following unregulated dumping of sewage from a growing urban population into the lake,

classic signs of eutrophication were observed, including blooms of Oscillatoria rubescens, reduced water

transparency, and very low nitrogen to phosphorus concentration ratios. The situation was studied

extensively by Dr. W. T. Edmondson, a professor at the University of Washington, who explained that

the changes in the lake were directly attributable to nutrient loading from sewage and wastewater

(Edmondson, 1991). Edmondson made these facts known not only to the scientific community but also

to the public and local government. The case resulted in the diversion of sewage away from the lake and

is a classic example of how scientific observations and understanding were used to shape public policy. 

The sewage diversion and subsequent lake recovery were a success, with current day water quality of the

Lake far exceeding that observed during the 1950's-60's.

The solution to this classic case was to divert wastewater from Lk. Washington to nearby Puget Sound, a

large inland sea linked to the Pacific Ocean via the Strait of Juan de Fuca. Studies by Dr. G. C.

Anderson, an oceanographer at the University of Washington, showed that phytoplankton in the Sound in

the vicinity of the outfall proposed to handle the diverted wastewater were limited by light and mixing,

not by nutrients. Thus, diversion of effluent from the Lake to the Sound was not "shifting the problem"

but rather was an ecologically sound solution. This understanding of both limnology and oceanography

laid the conceptual foundation for the formation of a large publicly funded agency (Municipality of

Metropolitan Seattle, or "Metro"). It was proposed that Metro would build a new outfall at West Point,

on the Main Basin of Puget Sound, divert the sewage from Lk. Washington to West Point, and thus

eliminate the nutrient enrichment problem. However, the mandate to create Metro to carry forward these

actions had to be approved in a public election first. Much controversy was associated with the process

chronicled well in Edmondson's book "The Uses of Ecology" (Edmondson, 1991). Among numerous

lines of objection, some public opinion maintained that Puget Sound's ecological health would be

destroyed, despite Anderson's observations. The proposed action required two election attempts before it

was approved in 1958. It is notable that the election passed before deterioration of Lk. Washington water

quality was serious; certainly the local conditions were not as serious as the symptoms seen in lakes in

Europe or the Midwest North America. However, deterioration of Lk. Washington conditions did

continue during the five years before the Metro diversion construction commenced. The persistent,

dense, and obnoxious populations of Oscillatoria galvanized public opinion that the Metro diversion was

necessary. Shortly after construction of the diversion, Lk. Washington water quality improved.

The notoriety of this event and the success of the outfall constructed at West Point to not exhibit

observable biological changes in Puget Sound resulted in widespread lore that "Puget Sound" cannot be

eutrophied because the marine waters are not sensitive to nutrient addition. As Anderson's observations

implied, the reason for the success of West Point outfall owes to the deep, well-mixed waters at the site

which are flushed with a residence time on the order of days. Density-driven stratification is minimal,

the phytoplankton are well mixed, and any depth gradients of oxygen and nutrients do not persist. The

outfall, at 71 m depth, diffuses effluent into water that has naturally high concentrations of nitrate and

this additional nitrogen burden is thought to not significantly contribute to phytoplankton nutrition. 
Nutrient Criteria—Estuarine and Coastal Water CS-43

Ammonium concentrations in excess of normal Puget Sound background levels are observed sporadically

(King County 2001) near the site and may be associated with the effluent. 

This example remains well-known, but it is important to note that not all of the reaches, bays, and inlets

of Puget Sound have the same characteristics of West Point and the Main Basin. Greater Puget Sound is

actually composed of several basins: South Puget Sound, Hood Canal, Whidbey Basin, the Main Basin,

and Admiralty Inlet with its adjoining waters with the Strait of Juan de Fuca. The first three of these

basins have considerable freshwater-induced stratification and are much less well-flushed than the Main

Basin. Residence times of these basins range weeks to months. A study funded by EPA to provide a

general review of the state of knowledge regarding nutrient-phytoplankton relations and quantify the

relative nutrient sensitivity of various areas in the Sound highlighted several areas where nutrients

became deplete and N:P ratios suggested nitrogen limitation (Rensel, 1991). Evidence from C-14 uptake

experiments have shown that enhancements of primary production over ambient rates due to added

nitrogen nutrient can be as high as 300% in South Hood Canal (Newton et al. 1995) and 83% in Budd

Inlet, located in South Puget Sound near Olympia (Newton et al. 1998). Based on environmental

attributes and human growth indicators, only a few places within Puget Sound were judged to be

currently exhibiting signs of eutrophication; however, numerous places, particularly in South Puget,

Hood Canal, and Whidbey basins, were assessed to be highly susceptible to future deterioration from

eutrophication (Bricker et al. 1999).

Perhaps the first place within Puget Sound to gain wide attention for nutrient enrichment effects was

southern Puget Sound, including Budd Inlet, Oakland Bay, Eld Inlet, Henderson Inlet, Case Inlet, and

Carr Inlet. The Washington State Department of Ecology sponsored two studies in the 1980s to evaluate

the acceptability of secondary-treated wastewater discharges to marine waters in southern Puget Sound

(URS, 1985, URS, 1986a). This work identified areas where new or expanded discharges were

unacceptable, based on the potential for eutrophication. A simple screening model based on effluent

dilution and flushing was developed to identify the most sensitive areas. 

Wastewater discharge into Budd Inlet was implicated in causing nuisance blooms of phytoplankton and

adding to low dissolved oxygen concentrations noted in the bottom waters at the head of the inlet (URS,

1986b). Studies sponsored by the Washington State Department of Ecology developed the first

numerical models to relate the loading of nitrogen to phytoplankton blooms and dissolved oxygen in

Budd Inlet. This work was the impetus for construction of advanced wastewater treatment systems to

remove dissolved inorganic nitrogen from municipal wastewater from the regional facility that discharges

to Budd Inlet. In 1994, the wastewater entering Budd Inlet was treated for nitrogen removal. This has

resulted in substantially lower ambient nitrogen concentrations (Eisner and Newton 1997). However,

regional growth continues and more capacity to process effluent is needed. Currently, the

Lacey-Olympia-Tumwater-Thurston County Wastewater Management Partnership (LOTT) is considering

numerous alternatives to meet this need, including one proposal to discharge more effluent in winter

when phytoplankton growth is minimal. LOTT undertook a large modeling and observational study: the

Budd Inlet Science Study, which will be used to make the final permitting decisions. 

Nutrient-induced increases to phytoplankton production with subsequent drawdown of bottom-water

oxygen has a unique character in Puget Sound relative to other North American systems because of the

influence of upwelled Pacific Ocean water. Upwelling favorable conditions off the Washington coast

lead to upwelling of deep ocean waters. These relatively low-oxygenated waters are transported in,

landwards, at depth through the Strait of Juan de Fuca, as the estuarine flow of Puget Sound waters flow

out, seawards, at the surface. The oceanic waters entering the Puget Sound system through Admiralty

Inlet can have oxygen concentrations as low as 5 mg/L, which then spread into the Main Basin and form

the bottom-water of the other basins as well. Seasonally low deepwater oxygen concentrations can be
CS-44 Nutrient Criteria—Estuarine and Coastal Water

found throughout Puget Sound (Newton et al., 1998; King County, 2001). Upwelling is mostly

favorable in late summer, when productivity-related oxygen deficits are also maximal. The additive

nature of human-caused eutrophication oxygen drawdown to this natural low oxygen quality, results in a

smaller margin of error before deleterious effects would be noted.

The success of discharging nutrients into marine waters via regional Puget Sound wastewater plants,

including the West Point facility, may be at a scale that has reached capacity. The Seattle metropolitan

area continues to grow and new Puget Sound regional wastewater facilities are needed. The King County

Council recently approved that a new facility will be required to meet growth demands and is projected

for completion in 2010. The new facility will also discharge into the Main Basin, to the north of the

West Point facility. King County is currently leading an extensive study to investigate impacts from

nutrient loading on the area. These studies include extensive water quality sampling, experiments on the

susceptibility of the waters to nutrient additions and modeling experiments to predict future impacts.

The story of nutrient control in Puget Sound continues to evolve. Ecologically sound decisions regarding

nutrient management are dependent on two variables: (1) the population producing the effluent, its size,

growth rate, and scale relative to the receiving waters; and (2) the sensitivity of the particular region of

the Sound where the effluent is to be discharged to nutrients. Unfortunately, both of these attributes are

highly variable within the Puget Sound regional area, making nutrient management decisions a challenge

and arguing for the utility of careful scientific studies in companion with planning efforts.
Nutrient Criteria—Estuarine and Coastal Waters A-1

APPENDIX A

CONDITIONS FOR BLOOM DEVELOPMENT:

INTERPLAY AMONG BIOGEOCHEMICAL, BIOLOGICAL, AND 

PHYSICAL PROCESSES

Overview

In temperate estuaries, the spring bloom typically is dominated by diatoms and occurs when freshwater

delivers adequate amounts of N, P, and Si and other nutrients to the system. In deep estuaries, the spring

freshwater inflow also provides for vertical density stratification where enough of the euphotic zone

resides near the surface to allow phytoplankton to achieve net biomass production (i.e., total

photosynthesis minus respiration is positive). During this time solar insolation and water temperature are

on the increase. For a bloom to develop, other conditions must be met. Physical flushing and local

dispersion of a water parcel must be less than the doubling time for cells. Biological grazing rate must

not be so large as to consume phytoplankton faster than cells’ doubling time. 

Biogeochemical Processes

Three aspects of N and P biogeochemistry help explain whether N or P dominates nutrient limitation in

estuaries (i.e., relative nitrogen fixation rates, denitrification, and sediment regeneration of P). The first

aspect involves evidence suggesting that nitrogen fixation is less effective in marine than in freshwater

systems in making up nitrogen deficits (Howarth 1988; Schindler 1974). This finding has major

implications for long-term coastal and open-ocean nutrient overenrichment, because N fixation is so

inefficient that any balance in the N versus P limitation occurs in terms of geological time.

The second aspect involves one of the greatest differences in nutrient biogeochemical cycles between

freshwater and marine systems. The superior capacity of freshwater versus marine systems to retain P in

sediments through interactions with iron has profound implications. Nearly all the P deposited in marine

sediments is remineralized annually (Caraco et al. 1990) and depends heavily on the sulfate

concentration, which can be used as a surrogate for salinity. Thus, P in freshwater sediments is bound

more tightly, and proportionally less is released back into the water column. Also, P release from marine

sediments is temperature dependent, and its maximum release during the summer helps explain the

tendency for increasing water column concentrations of P to occur during that season in many estuaries

(Nixon et al. 1980). Estuaries with a well-developed tidal freshwater zone might be expected to be more

P-limited than estuarine systems with small tidal fresh areas. 

More details on nutrient cycling in subtropical and tropical marine waters, systems much less studied

than northern temperate estuaries, are provided by Bianchi et al. (1999). The extensive coastal wetland

systems (e.g., marshes and mangroves) that border the Gulf of Mexico provide environments where

chemical transformations and storage of nutrients occur. Also, extensive seagrass meadows apparently

tie up inorganic N and P so that relatively less remains free in the water column. In general, DIN and

PO4 concentrations are much lower in northern Gulf river-dominated estuaries (e.g., Mobile and

Apalachicola Bays) than similar U.S. East Coast systems, presumably because of the lower point sources. 

Local groundwater sources are important and water quality managers should be aware of them. 

However, local estuarine point sources of N and P may alter nonpoint source patterns (e.g., paper mills

and wastewater treatment facilities) (Livingston 2001a).

Denitrification is the third aspect that plays a role in N limitation in estuaries and coastal waters (Nixon

et al. 1996). Denitrification is the process whereby nitrate is converted to gaseous N2 and N2O and

thereby made unavailable. Denitrification provides a sink for N in estuarine systems. Shelf waters

generally are too deep to provide enough sediment-water column contact for a quantitatively significant

magnitude of denitrification to occur. Bottom-water anoxia limits nitrification and hence denitrification
A-2 Nutrient Criteria—Estuarine and Coastal Waters

in the high-sulfide sediments where nitrification and denitrification are coupled (Jenkins and Kemp

1984). Knowledge of the magnitude of denitrification can help the water quality manager predict the

nutrient overenrichment response of an estuary, because N that is denitrified is largely unavailable to

support primary production.

Biological Processes

The relative importance of biological grazing should be assessed, because when a nutrient problem

occurs it is evidence that enrichment has exceeded the ability of the system to maintain a steady state in

net biomass production at pre-enrichment levels. For example, major changes in the biology of estuaries

in terms of particle filtering capacity (e.g., oysters: Newell 1988) and probably filter feeding finfish (e.g., 

menhaden) can modify phytoplankton primary production, although the quantitative effectiveness of such

cropping is scientifically unsettled. Nutrient overenrichment may drive marine waters toward smaller

algae and other microbes (Jonas 1992) where organic carbon flows more to the microbial loop (Hassett et

al. 1997), versus more direct flow to copepods and higher trophic levels (e.g., finfish). This shift in the

food web may be a significant factor in how estuaries “assimilate” increased nutrient inputs (Roelke et al.

1999; Roelke 2000). 

Physical Processes and Factors

Conceptual Framework

Smith (1984) argued that there is no inherent difference in nutrient limitation between lakes and the

ocean. Abundant evidence indicates that phytoplankton net primary production in north temperate lakes

tends to be P limited, and phytoplankton production in the ocean as a whole is potentially moderately P

limited but at higher P concentrations than lakes. This conclusion is supported by the observation that

TN:TP ratios of the surface ocean are usually well in excess of the Redfield ratio (Guildford and Hecky

2000). Local deviations have been detected. In contrast to lakes and oceans, estuaries and coastal shelf

waters tend to be N limited, with some exceptions. Water quality managers may question the reason for

this, as the three case studies described earlier, especially temperate estuaries and the coastal shelf,

appear to be N limited and not P limited. Such an understanding is basic to arguments about cause and

effect and also what ecosystem conditions drive a coastal ecosystem toward N limitation or P limitation. 

Smith provided an explanation that still has merit.

Smith posited that the apparent difference in limiting nutrient between lakes and oceans lies, in part, with

the relative rates for material exchange via physical processes of advection (i.e., transport of water and

associated constituents) and eddy diffusion (i.e., local transport of material against a concentration

gradient) and biogeochemical processes of N fixation and fixed N loss. The argument is based on field

experiments in marine embayments with little or no freshwater input, so advective transport of nutrients

from the land simplified nutrient budget development. Smith postulates that if the physical exchange

rates are long (e.g., open ocean), then the system would tend toward P limitation because biogeochemical

adjustment of the N:P availability ratio is short compared with long physical exchange rates (e.g., months

to a year or longer). In other words, nitrogen fixation would balance any losses of nitrogen associated

with phytoplankton sedimentation, but P has no atmospheric reservoir or biochemical mechanism for an

equivalent P fixation to occur. If the physical exchange rates are faster than biochemical rates (e.g.,

nitrogen fixation), then net ecosystem production (and by inference net phytoplankton biomass

production) of organic material may be N limited; if the biochemical rates are faster, then net production

will tend toward P limitation. The ratio of the residence time of the water to the biogeochemical turnover

rate indicates the degree to which the hydrodynamic processes dominate or modify estuarine ecosystems

(Day et al. 1989). This is an example of the importance of scaling critical processes (Harris 1986). 

Smith’s conceptual model should apply to estuaries and shelf waters. Smith’s data, analysis, and

synthesis and other empirical data support N limitation in estuaries and coastal shelf environments.
Nutrient Criteria—Estuarine and Coastal Waters A-3

Figure A-1. Scales of phytoplankton ecology. Horizontal and vertical scales

are determined by the respective diffusion coefficients Kh

 and KL. The time

scales for the algae are determined by the scales of growth (shaded band). The

processes of importance at each scale are noted. In the past it was always

assumed (wrongly) that physiological processes were at equilibrium and that

climatological variability could be ignored. Source: Harris 1986.

The concept of scale is another element of the conceptual framework. Physical processes that modify the

expression of bloom dynamics will best be detected at the ecological level (Figures A-1, A-2). In this

context, Geyer et al. (2000) cite many examples supporting the observation that “at virtually every

spatial scale, within every component of estuarine ecosystems, physical processes influence the

distribution and fate of chemicals (sic including nutrients) and organisms.” Physical processes are

involved in the delivery of nutrients to the biota in estuaries and coastal shelf waters, and also

fundamentally influence advective and dispersive processes that transport and retain dissolved and

particulate material, including nutrients and plankton in estuaries and on the coastal shelves. The roles of

physical processes influencing net biomass production of phytoplankton are explored in the main text in

more detail.
A-4 Nutrient Criteria—Estuarine and Coastal Waters

Figure A-2. The left panel shows the distribution of chlorophyll—an indicator of algal biomass—along the east

coast of the U.S. from Boston to South Carolina as measured from the ocean color satellite SeaWIFs. Note the

higher chlorphyll levels closer to shore, and the much higher levels in enclosed bays, such as Pamlico Sound

(latitude 35°) and Chesapeake Bay (mouth at 37° latitude). The above panel shows chlorophyll distributions within

Chesapeake Bay in more detail, as measured during a phytoplankton bloom. Both images were taken in April 1998. 

Source: Howarth et al. 2000.
Nutrient Criteria—Estuarine and Coastal Waters B-1

APPENDIX B

ADDITIONAL INFORMATION ON THE ROLE OF TEMPERATURE AND LIGHT 

ON ESTUARINE AND COASTAL MARINE PHYTOPLANKTON

Availability in temperature and light values can be plotted as hydroclimographs (e.g., polygons) to

picture their relative seasonal change around the coasts. A temperature increase will lower the density of

seawater and contribute to density stratification. In particular, rapid changes in water temperature

influence the rate of biological metabolic processes, including algal growth rates (Eppley 1972). Some

species exhibit various degrees of thermal adaptation when temperature changes are gradual.

Speculation suggests that if sea temperatures continue to rise as a function of the “greenhouse effect,”

estuarine biotic communities may change over the next several decades as they approach thermal limits. 

For example, although the seagrasses Halodule (shoalgrass) and Zostera (eelgrass) now overlap in Core

Sound, NC, a northward migration of Halodule and a retreat of Zostera may occur if water temperatures

rise faster than populations can adapt. Such relationships and their potential alteration probably can be

documented for other biotic groups along other coasts. For example, if temperature now limits regular

flowering of the seagrass Thalassia (turtle grass) along the northern Gulf of Mexico, then increased

flowering may occur if temperatures warm. Such conjectures notwithstanding, however, little

information is available to help assess the consequences of a potential interaction between an increased

temperature rise and increased nutrient supply on seagrasses.

Light has a fundamental role in aquatic primary production and is essential in the development of models

to estimate phytoplankton primary production (Behrenfeld and Falkowski 1997) and submerged aquatic

vegetation (Dennison et al. 1993). Many concepts in aquatic ecology are based on the light gradient

(Huisman 1999) (e.g., diel plankton vertical migration, benthic animals migrating out of sediments, depth

of euphotic zone, and mixing depth). Phytoplankton growth, nutrient relationships, light, and other

physical processes interact in a feedback system. Although light may be adequate and all other

requirements met for their formation, blooms may not form if dispersive processes are greater than algal

cell doubling time (Kierstead and Slobodkin 1953; Lucas et al. 1999). Tidal ranges greater than

approximately 2.0 m apparently disperse phytoplankton faster than cell doubling time, even if nutrient

conditions would be supportive of a bloom (Monbet 1992).

Both the vertical distribution of phytoplankton abundance and community composition are frequently

changing in the water column. Swimming through use of a flagellum, especially by dinoflagellates, and

changes in cell density through physiological mechanisms allow modest vertical mobility against weak

mixing forces. If the mixing depth is substantially greater than the euphotic zone depth (e.g., a depth

where approximately 1% of surface insolation occurs), then phytoplankton spend too much time in an

inadequate light environment and net primary production is limited (Figure B-1) (Huisman et al. 1999). 

The compensation depth is where water column phytoplankton photosynthesis and respiration are in

balance, and this often approximates the 1% insolation depth, or about two times the Secchi disc depth

(Parsons and Takahashi 1973). For example, in the lower Delaware Bay the upper-mixed layer often

corresponds to the bay bottom, the result of which is that phytoplankton spend too much time below the

compensation depth and hence low biomass production occurs (Pennock 1985). In systems where the

dominant support of the food web is derived from photosynthesis, net phytoplankton production must be

large enough to support the microbial loop (Azam et al. 1983) and higher trophic levels.
B-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure B-1. Diagram illustrating theoretical distribution of

phytoplankton production and phytoplankton respiration. After

Sverdrup (1953). Source: Mann and Lazier (1996).

In a homogeneous medium, light decreases exponentially with depth and can be represented by the

negative exponential equation:

Iz = Io e- kz

where Iz is light quantity at depth z, Io is the light quantity at the water surface, and k represents the

vertical light extinction coefficient; the extinction coefficient is more easily calculated as the base log10: 

0.434 kz = log I0 - log Iz). The light gradient also often extends longitudinally down estuaries, especially

those dominated by large volumes of sediment-bearing freshwater. In highly turbid estuaries, deepest

light penetration shifts toward the orange end of the spectrum (Champ et al. 1980). The euphotic zone

depth generally increases from the tidal head to the coast. Where turbidity is at its maximum level, a

localized sharp decrease in euphotic zone depth is typical (Flemer 1970; Pennock and Sharp 1986). 

Regions with the greatest turbidity typically are light-limited or almost so. In the turbid upper

Chesapeake Bay, riverine loading supplied nearly 90% of the particulate organic carbon, but in the

clearer waters of the mid-bay primary production dominated supply at 97% (Biggs and Flemer 1972); 

Smith and Kemp (1995) have updated their original estimate and suggest a lower percentage. However,

much of the allochthonous organic matter may not be biologically available. In waters of high humic

material content (e.g., Charlotte Harbor, FL), light attenuation can be severe (Dixon and Kirkpatrick

1999). The interaction between turbidity caused by humic materials and nonchlorophyll-bearing
Nutrient Criteria—Estuarine and Coastal Waters B-3

particulates complicates the direct application of the Secchi disc as a measure of nutrient overenrichment. 

Spectral radiometers can to some extent partition the various components of light extinction and are the

preferred tools. Modern algal pigment diagnostic tools (e.g., HPLC) can compare water quality

responses to varying nutrient and other pollutant inputs across various coastal system types (Jeffery et al.

1997). It is possible that turbidity may mask the impending development of undesirable algae.

The vertical extinction coefficient for estuaries shows wide seasonal variations. Values frequently

exceed 0.1 m-1). The extinction in open-ocean waters is often estimated by the relation E.C. = 1.7/

Secchi depth in meters. Holmes (1970) and Keefe et al. (1976) both arrived independently at a constant

of approximately 1.46 instead of 1.7 for turbid estuarine and nearshore coastal waters. Walker (1980)

suggested that the original Poole and Adkins (1929) Secchi disc constant on average gives results

approximately 17% too high and suggested a value of 1.45. These corrections should be made and, more

importantly, it is useful for the constant to be checked for each estuary. For more quantitative work, a

quantum light meter that measures over a spectral range of 400-700 nM is preferred. Commercial

products (e.g., www.licor.com) are now available that can measure the spectral photon flux over a range

of interest to aquatic scientists.

The extinction coefficient can be broken down into several components. The total light attentuation, KT

= Kw + Kc + Kd + Kp (Lorenzen 1972; Kirk 1983; Bledsoe and Phlips 2000; Koenings and Edmondson

1991) can be resolved for the effects of water, chlorophyll a, dissolved substances, and nonalgal

particulate matter. In many estuaries, Kd may contribute between 5% and 50% of the KT. The Kw

usually can be ignored because it is such a minor component. In blackwater estuaries receiving high

loads of humic materials the Kd may dominant KT. In systems such as the “turbidity maximum zone” in

upper Chesapeake Bay, Kp may be the dominant component. The EPA Chesapeake Bay Program has

sponsored research to calibrate KT components applicable to SAV beds (www.chesapeakebay.net; search

the publications database for “Chesapeake Bay Submerged Aquatic Water Quality and Habitat-Based

Requirements and Restorations Goals: A Second Technical Synthesis.”

Among some coastal ecosystems, light (i.e., mean photic depth) and nutrient loading appear to be equally

good predictors of phytoplankton primary production (see Figure 1b in Cloern 1999). This observation

strengthens the proposition that phytoplankton production in these systems can be limited by other

resources and processes in addition to nutrient loading. Pennock and Sharp (1994) suggest that the

Delaware River Estuary functions analogously as a chemostat during the summer. They point out that

high N supplies from upstream advect continually through the brackish-water region into the lower

estuary, where high primary production occurs from remineralized N and the advected supply and the

phytoplankton biomass is limited by grazing. Their conclusion is especially significant because evidence

also suggests that bioassay experiments that isolate the water may lead to misidentification of nutrient

limitation. The flushing component in the bay provides the physical analogue to a chemostat, where the

nutrient supply and adequate light support high phytoplankton biomass production and grazing and

flushing maintain a potentially steady-state phytoplankton biomass, with fluctuations due primarily to

physical forcing factors.
Nutrient Criteria—Estuarine and Coastal Waters C-1

APPENDIX C

ADDITIONAL INFORMATION ON FLUSHING IN ESTUARIES

by Edward H. Dettmann, U.S. EPA, Office of Research and Development, National 

Health and Environmental Effects Research Laboratory, Atlantic Ecology Division

A variety of terms such as residence time, flushing time, transit time, turnover time, and age are used to

describe time scales for transport and removal of materials that enter waterbodies. Use of these terms in

the literature is often inconsistent and sometimes imprecise, so that care must be exercised to determine

the meaning of terms being used, to avoid misinterpretation or incorrect comparisons of data. This

appendix contains definitions of terms and discussions of factors that affect flushing and empirical and

modeling methods for estimating residence times in estuaries. An understanding of residence times is

especially important when estimating not only system responses to nutrients, but also the lag phase

between management and system improvements.

Definitions of Residence Times

The freshwater residence time (Jfw) is the average amount of time that freshwater, or a conservative tracer

introduced with freshwater inputs, resides in the estuary before exiting. It is the mean transit time for a

molecule of water or conservative tracer to progress from the point of input to the seaward boundary.

This is the definition of residence time most often given in the literature, and is generally the most useful

for analysis of eutrophication in estuaries, as most nutrients are usually introduced with freshwater. 

Another commonly used concept is that of the mean amount of time required for water (or a

homogeneously-distributed conservative tracer) that is in the estuary at a given time (regardless of

source) to leave the estuary. This is here called “estuary residence time”(Je), although Takeoka (1984)

and Zimmerman (1976) use the term “residence time.”

The values of Jfw and Je may differ. Consider the case of an elongated unstratified estuary, with the

seaward boundary at one end and a river as the single source of freshwater located at the opposite end. In

this case, the travel distance from the river mouth to the seaward boundary is longer than the average

travel distance from other parts of the estuary, so that Jfw > Je. Conversely, if the river enters the estuary

near the seaward boundary, Je may exceed Jfw. For cases intermediate between these two, the difference

between these two residence times becomes smaller, and the times may be equal. Multiple freshwater

inputs and other factors such as stratification and estuary shape may complicate matters, but the same

principles apply.

These concepts, the related concepts of turnover time and age, and the relationships among these

measures are discussed in detail by Bolin and Rodhe (1973) and Zimmerman (1976), and summarized by

Takeoka (1984). Miller and McPherson (1991) discuss another concept, “pulse residence time,” for water

or a conservative constituent introduced as a pulse, that may be useful in some circumstances.

Factors That Influence Residence Time

Water residence times in estuaries are influenced by any factor that affects water movement, including

freshwater inflow rates, tides, wind, mixing, stratification, and system topography. Because many of

these factors are variable, residence times are also not static. This variability requires attention to the

appropriate time interval over which the residence time should be expressed, and the representativeness

of the conditions under which a given measurement is made. A long-term (seasonal or annual) average

residence time is often most appropriate for analysis of the effects of nutrients.
C-2 Nutrient Criteria—Estuarine and Coastal Waters

Freshwater Forcing

Residence time may be quite sensitive to freshwater inflow rate, with larger flow rates associated with

smaller residence times. This is illustrated by three Rhode Island estuaries. For Narragansett Bay, Jfw

varies between approximately 10 and 40 days; Jfw at the long-term mean freshwater inflow rate is 26 days

(Pilson 1985). For the Seekonk River, Jfw varies between 0.4 and 2.9 days, with a value of 1.2 days for

the mean freshwater inflow rate, and for the Providence River, Jfw varies between 0.7 and 6.3 days, and is

2.5 days at average inflow (Asselin and Spaulding 1993).

Tidal Forcing

Tides can be a major factor controlling estuary-ocean exchange of water and therefore water residence

time. Important factors are the tidal range, tidal frequency (diurnal vs. semidiurnal), and estuary depth.

Tidal ranges in U.S. coastal waters range from centimeters to more than 5.5 m. For a given estuary,

residence times may vary over the spring-neap tide cycle.

Wind Forcing

Wind may substantially influence estuarine circulation, and therefore water residence time. In a study of

two small shallow estuaries, Geyer (1997) found that wind direction (offshore vs. onshore) had a

substantial effect on salinity structure and nontidal flow in one estuary, and Jfw varied by a factor of

approximately 3 (from 0.8 to 2.7 days) in response to the sea-breeze cycle. Measurements in a nearby

and similarly oriented estuary during the same time period found that whereas wind direction influenced

salinity distributions, a constriction at the estuary mouth limited estuary-offshore exchange, so that there

was no significant relationship between residence time and wind stress. Estuaries bordering the Gulf of

Mexico are shallow, highly susceptible to wind forcing, and considered meteorologically dominated

(Solis and Powell 1999; Ward 1980).

Determination of Residence Time

A number of empirical and computational methods are used to measure or estimate water residence

times.

Empirical Measurements

Empirical measurements of water residence time depend on measurement of the distribution or dynamics

of tracers (generally freshwater or introduced dye) in estuaries. 

Bolin and Rodhe (1973) show that the transit time of a tracer through a reservoir is given by the turnover

time (Jo) , i.e. 

where Mo is the total mass of a constituent in the reservoir and Fo is the total flux through the 

reservoir. This is the basis for the freshwater replacement method for calculating the mean freshwater

residence time in an estuary. The Jfw is calculated as

where Vfw is the volume of freshwater in the estuary and Qfw is the input rate of freshwater. This ratio

gives the time required for the inflowing freshwater to replace the freshwater already in the estuary. The

volume of freshwater in the estuary is calculated as the amount of freshwater that must be mixed with
Nutrient Criteria—Estuarine and Coastal Waters C-3

seawater having salinity (Ss) equal to that entering at the seaward boundary to yield a volume Ve equal to

that of the estuary, with salinity equal to the mean salinity of the estuary (Se). Vfw may be calculated as

Examples of this method are given by Pilson (1985), Geyer (1997), and Solis and Powell (1999).

Fluorescent dye (usually Rhodamine WT) is also used to determine water residence times. To measure

freshwater residence time, dye is introduced continuously into the inflowing freshwater at a rate

proportional to the freshwater flow rate. Dye concentrations are surveyed at (high or low) slack tide. 

Dye input is terminated when the mean concentration of dye in the estuary reaches equilibrium

(approximately 3 times the freshwater residence time). Further periodic surveys are conducted at the

same tide phase to monitor the dye content of the estuary. The average concentration of dye is usually

found to follow a decreasing exponential with time,

where Co is the initial average concentration of dye in the estuary and C(t) is the concentration at time t. 

This function is fit to the data, and the mean residence time of dye in the reservoir, a surrogate for the

mean freshwater residence time, is then Jfw = 1/k. Alternatively, as the concentration one residence time

after termination of dye input is C = Co e-1, the time required to attain this concentration is sometimes

taken as the residence time. The estuary residence time may be measured similarly. In this case, the dye

is distributed as uniformly as possible throughout the estuary in a rapid application, and the change in

dye content is monitored as above. The residence time determined from the rate of change of tracer

concentration is sometimes called the e-folding time. Because of cost and logistical considerations, dye

studies are most often done on relatively small estuaries. Examples of such studies are described by 

Callaway (1981), Dettmann et al. (1989), and Geyer et al. (1997).

Models

A wide range of models, ranging from simple to complex, has been used to calculate water residence

time in estuaries. The simplest of these is the tidal prism model, which estimates residence time (in

number of tidal cycles) as the ratio V/P, where V is the estuary volume (usually expressed at high tide)

and P is the volume of the tidal prism. The model is based on the assumption that water entering the

estuary on the flood tide is thoroughly mixed throughout the estuary within a tidal cycle. Most estuaries

do not meet this requirement and, for all but very small estuaries, this model generally underestimates

estuary and freshwater residence time, sometimes by manyfold.

Ketchum (1951) modified this simple model by dividing the estuary into segments, each having a length

that corresponds to the local tidal excursion. Water exchanges between adjacent segments during each

tidal cycle are calculated, and complete mixing is assumed to occur only within each segment. The

segmented tidal prism model requires data for freshwater inputs, tidal range, and estuary topography. The

simple and segmented versions of the tidal prism model are reviewed by Pritchard (1952) and Dyer

(1973), and modified by Dyer and Taylor (1973). For most (but not all) estuaries to which the segmented

model has been applied, it has given good results (Dyer 1973). Beyond permitting calculation of

residence time, segmented tidal prism models permit calculation of the distributions of fresh- and
C-4 Nutrient Criteria—Estuarine and Coastal Waters

saltwater and other water quality constituents along the estuary, but do not address the effects of wind on

flushing.

Another approach to calculating residence times is the box model. Box models also segment the estuary,

assume complete mixing within segments, and calculate diffusive and advective exchanges between

adjacent segments. These models require data for estuary topography and for freshwater inflows and

salinity distribution in the estuary for each set of conditions to which the model is applied. As actual

salinity data are used, the effects of wind and tide are implicitly taken into account. Applications of such

models are described by Brown and Arellano (1980), Dettmann et al. (1992), Hagy et al. (2000), and

Miller and McPherson (1991). Such models also allow calculation of advective and diffusive exchanges

among segments and spatial distribution of water quality constituents.

Numerical computer models of hydrodynamics and constituent transport are still more complex. These

models are used to simulate movement of water and water quality constituents at fine spatial resolution,

and may also be used to calculate freshwater or estuary residence times. These models generally allow

consideration of wind stress as well as tidal and freshwater forcing. A few examples of such models are

described by Brooks et al. (1999), Signell (1992), and Signell and Butman (1992).
Nutrient Criteria—Estuarine and Coastal Waters D-1

APPENDIX D 

NOAA SCHEME FOR DETERMINING ESTUARINE SUSCEPTIBILITY

(SOURCE: BRICKER ET AL. 1999)

The following provides an overview of the NOAA scheme for determining the overall human potential

for causing nutrient enrichment of estuaries.

Dilution and Flushing Potential

The length of time that nutrients spend in an estuary potentially affects their opportunity to contribute to

overenrichment. The time is a function of dilution potential and flushing rate. The analysis uses

physical and hydrologic data to define separately (1) a dilution rating and (2) a flushing rating. In both

cases, the higher rating, the greater the capacity to dilute or flush nutrient loads (see Tables D-1 and 

D-2).

Figure D-1 combines dilution potential and flushing potential. By combining dilution and flushing

components, an export potential (EXP) is determined. Estuaries in the upper left portion of the matrix

generally have a high EXP that suggests an ability to dilute and flush nutrient loads. Estuaries in the

lower right portion of the matrix have the opposite capacity, making them more susceptible to nutrient

input.

Nutrient Inputs

NOAA used the USGS Sparrow Model (spatially referenced regressions of contaminant transport on

watershed attributes) and other information to estimate nutrient loads and measure nitrogen pressure on

estuaries. 

Determination of Overall Human-based Nutrient Pressure

A matrix was used to compare the susceptibility to nutrient retention and the level of N inputs to rank the

overall expression of human influence on eutrophic conditions in the estuary (Figure D-1). Experts at the

National Assessment Workshop reviewed and, where appropriate, modified the assessments based on

higher quality data available for some estuaries; expert knowledge also played a role.
D-2 Nutrient Criteria—Estuarine and Coastal Waters

Table D-1. Decision rules for dilution potential

Type IF: vertical stratification THEN: dilution volume IF: dilution value Dilution potential No. estuaries

A Vertically homogenous

•all year

•throughout estuary

1/VOLestuary 10!13

10!12

High 30

B Minor vertical stratification

•navigation channels

•upper estuary

1/VOLestuary 10!11 Moderate 63

C Vertically stratified

•most of year

•most of estuary

1/VOLfwf

(fwf=freshwater fraction)

10!10

10!09

Low 45

Note: This analysis assumes that a larger portion of the water column is potentially available to dilute nutrient loads in a

vertically homogeneous estuary than in a vertically stratified system. The assumption is that for stratified systems, nutrients are

most often retained in the upper portion (freshwater fraction) of the water column. In contrast, downward transport (more

complete mixing) is likely in vertically homogeneous systems. Type B estuaries are generally vertically homogeneous, although

stratification is observed (confined) in narrow navigation channels or the extreme upper reaches of an estuary. In this case,

nutrients are assumed to be diluted throughout the entire water column. 

Source: Bricker et al. 1999.

Table D-2. Decision rules for flushing potential 

Type Tide range (ft) Freshwater inflow/estuary volume Flushing potential No. estuaries

1 Macro (>6) and Large or moderate (1000 to 10!02) High 12

2 Macro (>6) and Small (10!03, 10!04) Moderate 21

3 Meso (>2.5) and Large (1000, 10!01) High 15

4 Meso (>2.5) and Moderate (10!02) Moderate 16

5 Meso (>2.5) and Small (10!03, 10!04) Low 26

6 Micro (<2.5) and Large (1000, 10!01) High 4

7 Micro (<2.5) and Moderate (10!02) Moderate 13

8 Micro (<2.5) and Small (10!03, 10!04) Low 31

Note: This analysis assumes that a greater capacity to flush nutrient loads exists for estuaries that have large tide and freshwater

influences.

Source: Bricker et al. 1999.
Nutrient Criteria—Estuarine and Coastal Waters D-3

Figure D-1. Estuarine Export Potential Susceptibility.
Nutrient Criteria—Estuarine and Coastal Waters E-1

APPENDIX E

COMPARATIVE SYSTEMS EMPIRICAL MODELING APPROACH:

THE EMPIRICAL REGRESSION METHOD TO DETERMINE NUTRIENT 

LOAD-ECOLOGICAL RESPONSE RELATIONSHIPS FOR 

ESTUARINE AND COASTAL WATERS

By James Lattimer, USEPA, Office of Research and Development, National 

Health and Environmental Effects Research Laboratory, Atlantic Ecology Division

The empirical regression method can be used to determine the response of estuarine systems to nutrient

loading. This method can take two forms: single system and comparative-systems approaches. 

In single system approaches, ecological responses and nutrient concentrations are measured over time in

a single system. This allows for the development of load-response models for the individual systems that

are robust because many of the controlling factors, such as, for example, physiographic setting, primary

productivity base, and hypsography, are relatively constant. This approach has been used to develop

models of the response of primary productivity to nitrogen load (Figures E-1a,b) (Boynton et al. 1995).

However, this approach is only applicable to the system in which it was developed and thus is not

considered to be very useful in the development of widely applicable nutrient load-response

relationships. Single system empirical studies, however, may be useful in providing data on processes

useful for numerical model development.

The alternative approach, using the space for time paradigm (Picket,1988), posits that relationships

between nutrient inputs and ecologically meaningful estuarine responses, using multiple systems, have

predictive capability, at least for the category of systems used in the model development. This allows for

a wide range in nutrient loading and estuarine types to be included. The comparative-systems empirical

approach has been used to determine, for example, relationships between nutrient inputs and fish yields

(Lee and Jones 1981; Nixon 1992), benthic biomass, production and abundances (Josefson and

Rasmussen 2000), summer ammonia flux (Boynton et al., 1995), chlorophyll-a concentration (Boynton et

al. 1996; Boynton and Kemp 2000; Monbet 1992), primary productivity (Nixon et al., 1996), and the

dominant source of primary productivity (Nixon et al. in press). In many of these cases, important

environmental factors such as flushing time and depth, are used to normalize the nutrient loading in a

similar way as Vollenweider (Vollenweider 1976) to yield more precise relationships. 

The comparative-systems empirical approach has been successfully used in a regulatory framework to

develop total maximum loads for 30 subestuaries within Buzzards Bay in Massachusetts (Costa et al.

1999). Using a citizens’ monitoring network, many important water quality variables were measured

during summer sampling periods over a five to seven year period. Nitrogen inputs were estimated from a

land-use model modified from Valiela (Valiela et al., 1997) augmented with literature data on point

source and atmospheric inputs. In this study, nitrogen load-response relationships were derived for

nitrogen concentration, chlorophyll-a, secchi depth, dissolved oxygen concentration, and eelgrass habitat

ratio. Nitrogen loadings were normalized to account for volume and flushing time of each of the systems

to improve the precision of the empirical models. An example of the types of relationships determined

using this approach are given in Figure E-2. The regional estuarine management program used this

method to adopt total maximum annual loads (TMALs) for nitrogen. Specifically, “...the nitrogen

management strategy represents a linking of estimates of watershed nitrogen loading...to empirical

observations of ecosystem response in a wide variety of Buzzards Bay embayments” (Costa 2000).

The comparative-systems empirical approach does not explicitly consider the processes that produce the

observed phenomena; however, factors known to affect behavior are determined in order to reduce the
E-2 Nutrient Criteria—Estuarine and Coastal Waters

Figure E-1a. Scatter plots showing correlation between nitrogen loading, expressed

using the volume Vollenweider-term flushing scale, and 92-98 mean +/- std. errors of the

Alternate Eutrophication Index scoring (without oxygen scores). 

uncertainty in the models. The smaller the variance in the load-response relationship, the more

compelling the association. One important question for management is to determine what level of

variance is sufficient to convince stakeholders to accept nutrient limits and associated monetary

expenditures. 

This approach is largely based on statistical associations and is therefore restricted to prediction within

the class of systems used in the model development. Applicability improves, however, by the inclusion

of systems that encompass a wide range in loading and responses. The comparative-systems empirical

approach allows for the direct measurement of important endpoints (e.g., hypoxia, SAV loss, biomass)

obtained in the environment. The endpoints are what are important to the general public. So by

providing a mathematical relationships between the stressor and important endpoints, managers can

convince the public of the importance of regulatory action.
Nutrient Criteria—Estuarine and Coastal Waters E-3

Figure E-1b. Ratio of eelgrass habitat area to potential habitat area versus nitrogen

loading, expressed using the volume Vollenweider-term flushing scale (from Costa et al.,

1999).
E-4 Nutrient Criteria—Estuarine and Coastal Waters

Figure E-2. Plot of annual TN loading rates versus phytoplankton primary production rates at a

single station in the Chesapeake Bay from 1971-76 and 1985-90. Source: Boynton et al., 1995.
Nutrient Criteria—Estuarine and Coastal Waters F-1

APPENDIX F

SELECTED THEORETICAL APPROACHES TO CLASSIFICATION OF 

ESTUARIES AND COASTAL WATERS

Several theoretical schemes have been presented that may facilitate classification of estuaries and coastal

waters if the need arises for more theoretical approaches. In most cases, these schemes have potential

future value but are not likely to be immediately useful. The information-thermodynamic approach

offered by Ulanowicz would seem to provide useful insights to development of an eutrophication index

and may be worth an earlier consideration.

Functional Attributes

Odum and Copeland (1974) proposed a classification scheme based on the idea that an ecosystem is a

balance between energies that build structure and order, or ordering energies, and energies that cause

loss of structure and order, or disordering energies. Although this approach is of theoretical interest, the

data to apply it are still largely unavailable. There are conceptual difficulties. Energy sources and

stresses are not always mutually exclusive categories. Day et al. (1989) give several examples of cases

where moving water can be either an energy source or a stress, depending on the situation. Moderate

currents are a source of energy for seagrass meadows, because they transport organic matter and

inorganic nutrients to beds and remove metabolic wastes. If currents are too strong, however, they can

disrupt beds and act as disordering energies. 

Theoretic Indices

This approach is data intensive and has potential as appropriate estuary data are generated. Ulanowicz

(1986, 1997) described an approach based on flow analysis and information theory. The idea is that an

ecosystem can be characterized in terms of growth and organization. Growth is defined as an increase in

system activity or total system throughput (analogous to total system energy flow). Organization is

equated with the mutual information inherent in the trophic flow structure. To apply this approach, one

would need to obtain energy flow values (or organic carbon-based equivalents) among trophic

compartments. Following Ulanowicz’s formulation, one could develop an information index of

eutrophication. Compiling such would entail an assessment of the growth and organization status of a

eutrophic or nutrient-enriched system compared with a current reference system, a minimally impaired

system, or an estimate of pre-eutrophication values from historical data. The ratio of growth and

organization of a nutrient-enriched system to its reference value would express the degree of impairment. 

Such information could help guide restoration priorities. The downside is that the approach requires the

availability of a rich ecological database.
Nutrient Criteria—Estuarine and Coastal Waters G-1

APPENDIX G

EXAMPLES OF NUTRIENT CONCENTRATION RANGES AND RELATED

HYDROGRAPHIC DATA FOR SELECTED ESTUARIES AND COASTAL WATERS IN

THE CONTIGUOUS STATES OF THE UNITED STATES

Atlantic Coast Systems

BOSTON HARBOR, MA (Source: Kelly 1998)

WAQUOIT BAY, MA (Source: Valiela et al. 1992) 

SEEKONK-PROVIDENCE RIVER REGION OF NARRAGANSETT BAY, RI 

(Source: Doering et al. 1990)

WESTERN LONG ISLAND SOUND AND HUDSON-RARITAN ESTUARY, NY/NJ 

(Source: O’Shea and Brosnan 2000)

HUDSON RIVER, NY (Source: Lampman et al. 1999)

DELAWARE RIVER ESTUARY (Source: Lebo and Sharp 1993)

MARYLAND COASTAL BAYS (Source: Boynton et al. 1996)

CHESAPEAKE BAY, MD/VA (Source: Magnien et al. 1992) 

CHESAPEAKE BAY (Harding and Perry 1997)

YORK RIVER ESTUARY, CHESAPEAKE BAY, VA (Source: Sin et al. 1999)

NEUSE RIVER ESTUARY, NC (Source: Rudek et al. 1991)

CAPE FEAR RIVER ESTUARY, NC (Source: Mallin et al. 1999)

COASTAL GEORGIA (Source: Hopkinson and Wetzel 1982)

Gulf of Mexico Estuaries

GULF OF MEXICO CASE STUDIES (Source: Bianchi et al. 1999)

GALVESTON BAY, TX (Source: Santschi 1995)

Pacific Coast Systems

EMBAYMENT OF PUGET SOUND, WA (Source: Bernhard and Peele 1997) 
G-2 Nutrient Criteria—Estuarine and Coastal Waters

The study area (N latitude and W longitude) in Boston Harbor

and western Massachusetts Bay. The boundary between the

Harbor and the Bay is defined by the solid line from Deer

Island to Hull combined with the dashed lines showing the 2

high-resolution transects, the spatial limits of data for box

modeling are depicted. Dots position some watercolumn

monitoring stations that were sampled during 1994. Stations

prefixed with ‘N’ surround the future offshore outfall, which is

centered between Stns N20P and N16P about 15 km from

Deer Island. Data from 2 lines of Bay staions (N01P to N10P

and N04P to N07P, each representing about 10 km distance)

were used to provide approximate concentrations for the future

tidal source region for the Harbor.

Boston Harbor
Nutrient Criteria—Estuarine and Coastal Waters G-3

Bay (C0) and Harbor (C1) concentration data used in calculating Harbor-Bay exchange. BA: beam attenuation from

transmissometer readings; nd: not determined.
G-4 Nutrient Criteria—Estuarine and Coastal Waters

Time course of light, temperature,

oxygen, ammonium, phosphate, and

chlorophyll in water of Waquoit Bay,

summer 1988. Modified from Costa et al.

(in press).

Waquoit Bay
Nutrient Criteria—Estuarine and Coastal Waters G-5

Vertical profiles of oxygen in Childs River during dawn and midafternoon

of a sunny day (circles), and during afternoon after three cloudy days in a

row. Data from C. D’Avanzo.
G-6 Nutrient Criteria—Estuarine and Coastal Waters

Station locations, Solid lines delimit boxes used

in modeling effort. BR=Blackstone River, BV=

Blackstone Valley Sewage Treatment Plant

(STP), TM= Ten Mile River, MR Moshassuck

River, WR=Woonasquatucket River, FP=Field’s

Point STP, EP=East Providence STP, PR=

Pawtuxet River

Seekonk—Providence River Region,

Narragansett Bay
Nutrient Criteria—Estuarine and Coastal Waters G-7

Mean and range of salinity and nutrient concentrations observed in

surface waters (depth=1.0m) during the six cruises, versus distance

from the Main Street Bridge, Pawtucket, Rhode Island at the head

of the Seekonk River.
G-8 Nutrient Criteria—Estuarine and Coastal Waters

Mean and range of salinity and nutrient concentrations observed in bottom

waters (1.0 m from bottom) during the six cruises, versus distance from the

Main Street Bridge, Pawtucket, Rhode Island at the head of the Seekonk

River.
Nutrient Criteria—Estuarine and Coastal Waters G-9

Location of New York City Department of Environment Protection’s Harbor Survey Water Quality

Monitoring Stations. Also depicted are the 14 New York City Water Pollution Control Plants and the 8

New York City sludge dewatering facilities (shaded boxes). Sandy Hook, lower center, is located at

approximately 40°30'N, 74°W.

Western Long Island Sound

Hudson-Raritan Estuary

LOCATION OF WPCP’s AND SAMPLING STATIONS IN NY HARBOR
G-10 Nutrient Criteria—Estuarine and Coastal Waters

Key water quality indicators for summer (June–September) of 1999. Depicted are: bottom minimum

dissolved oxygen (upper left) in mg l!1; summer average dissolved inorganic nitrogen

[NH3!N+(NO3+NO2)!N] (upper right) in mg l!1; summer average Secchi transparency (lower left) in m;

summer average total phosphorus (lower right) in mg l!1.
Nutrient Criteria—Estuarine and Coastal Waters G-11

1989–1999 summertime (June–September) average ambient dissolved inorganic nitrogen [dissolved

ammonium-nitrogen (NH4-N) and dissolved nitrate- and nitrite-nitrogen (NO3)-N] concentrations (mg l!1 for

western Long Island Sound, Raritan Bay, and Jamaica Bay stations. Trends in summertime average

concentrations are noted where significant (p < 0.05).
G-12 Nutrient Criteria—Estuarine and Coastal Waters

1989–1999 summertime (June–September) average ambient total phosphorus (TP) and dissolved

orthophoshpate (PO4-P) concentrations (mg l-1) for western Long Island Sound, Raritan Bay, and Jamaica

Bay stations. Trends in summertime average concentrations are noted where significant (p < 0.05).
Nutrient Criteria—Estuarine and Coastal Waters G-13

Map of the tidal freshwater Hudson River, its watershed

and major tributaries. The inset shows the location of the

watershed of the watershed of the Hudson, primarily in

eastern New York State. The main panel shows the

watershed of the Hudson. The upper (non-tidal) tributaries

are in white and their surrounding watershed is shaded. 

The remaining, unshaded, section of the watershed

delivers water to the tidal Hudson River (heavy black line)

primarily through 7 tributaries. These tributaries (creeks)

are labeled: K for Kinderhook; Cfor Catskill; RJ for

Roeloff Janson Kill; E for Esopis; R/W for

Rondout/Wallkill; WP for Wappingers; Cr for Croton. 

The first 5 of these tributaries enter within the budgeting

reach (labeled) between km 225 and 125. The section of

the river that was sampled during our transect sampling

(labeled sampling stretch) is also shown.

Hudson River, NY
G-14 Nutrient Criteria—Estuarine and Coastal Waters

Box and whisker plots for forms of N and P for the

entire data set of the Hudson. The data include all

samples taken from January 1992 to December 1996

and combine seasonal and spatial variation. Shown are

the medians, upper and lower quartiles and 90%

inclusion lines. The number of samples for each

analysis is labeled near the boxes.
Nutrient Criteria—Estuarine and Coastal Waters G-15

Spatial variation in selected variables for a

representative transect taken in early September

1996. Transects run km 240 (Albany

downstream to km 40. For each variable

samples were taken every 2 to 4 km. Units are,

conductivity—(note log scale); pH—normal pH

units; chlorophyll a— 1-1; turbidity— NTU; N

and P—μM. For both N and P the per line

represents total N and toal P. “Other N” is TN

minus DIN (NH4 plus NO3); For P, “Other P” is

TP minus phosphate.
G-16 Nutrient Criteria—Estuarine and Coastal Waters

The Delaware Estuary with insert showing relation to

the mid-Atlantic coast of the United States. Stations

(1-9, 11-26) were located along the main axis of the

estuary, every 5-15 km. Station 10 was located in the

Schuylkill River (not shown). The river is tidal from

the mouth of the bay up to the fall line at Trenton,

New Jersey (240 km). Salinity increases from 0.1%

at station 14 (130 km) to 30% at the bay mouth. The

majority of TP inputs to the river are clustered

between stations 6-9.

Delaware River Estuary
Nutrient Criteria—Estuarine and Coastal Waters G-17

Figure 16Average biological, chemical, and physical parameters for Delaware Estuary 1986-1988. Values

are the nonweighted mean of 17 to 24 discrete samples taken at each location (Lebo et al. 1990). Data are

shown for locations near the Delaware River (station 2), Philadelphia (station 7), beginning of the salinity

gradient (station 14), turbidity maximum (station 16), downstream of the turbidity maximum (station 20), and

in the lower bay (station 23). In addition, average relative standard deviation (RSD) for all stations is shown. 

The average standard deviation for station location and salinity was 1.1 km and 2.1%, respectively.
G-18 Nutrient Criteria—Estuarine and Coastal Waters

Map of the Maryland coastal bays complex indicating the boundaries of the watershed and a

subsystems for which nitrogen inputs were estimated.

Maryland Coastal Bays
Nutrient Criteria—Estuarine and Coastal Waters G-19

Summer average chlorophyll a concentrations for representative regions of the Maryland coastal bays

based on samples collected during 1975, 1983, and 1991. Data are from Fang et al. (1977a,b), Maryland

Department of Health and Mental Hygiene (1985), and National Park Service (1991).
G-20 Nutrient Criteria—Estuarine and Coastal Waters

Map of Chesapeake Bay region showing tributaries, sampling sites, and major wastewater

treatment plants for systems examined or referenced in this paper. The major wastewater

treatment plants are as follows: 1) Western Branch, 2) Arlington, 3) Blue Plains, 4)

Alexandria, 5) Piscataway, 6) Little Hunting Creek, and 7) Lower Potomac.

Chesapeake Bay

Magnien et al.
Nutrient Criteria—Estuarine and Coastal Waters G-21

Salinity dilution plots of nitrate plus nitrite,

ammonium, total nitrogen, orthophosphate, and total

phosphorus for surface samples at all Chesapeake

Bay Mainstream stations. Each point represents a

single station at which median values were calculated

for each seasonal time frame (see Fig. 2) over the

entire 195-1989 period. All stations identified in Fig.

1 were plotted in longitudinal order starting at the

head of the estuary. The circled point for each

seasonal plot is the upper mesohaline plankton station

at the Annapolis Bay Bridge.
G-22 Nutrient Criteria—Estuarine and Coastal Waters

Concentrations of ammonium (μM) in Back River, and ammonium concentrations, primary production

estimates (mg C m-2 d-1) and temperature (BC) for the Patapsco River(Baltimore Harbor). Values presented

are surface mixed layer medians by month from continuous monitoring at a frequency of one to two times per

month over the period 1985-1989. Station locations are centrally located in each system (see Fig. 1)

Month

Back River Patapsco River

NH4

+ NH4+ Primary Production Temperature

January 257.0 8.9 299 3.2

February 414.0 48.2 1,384 3.0

March 112.4 24.3 640 6.2

April 439.0 33.6 663 12.3

May 151.7 11.7 1,520 16.8

June 35.7 9.7 1,840 24.3

July 82.1 13.0 2,263 27.2

August 87.5 30.3 1,950 26.6 

September 107.1 18.8 1,744 23.8

October 103.5 21.3 953 17.7

November 232.0 15.4 751 11.3

December 317.7 13.1 930 6.5
Nutrient Criteria—Estuarine and Coastal Waters G-23

Chesapeake Bay showing locations of the 6 regions.

Chesapeake Bay

Harding and Perry
G-24 Nutrient Criteria—Estuarine and Coastal Waters

Regions in the Chesapeake Bay by latitude and geographic location.

Region Latitude range Geographic location

I 36.95-37.40NN Mouth of Bay to Mobjack Bay

II 37.41-37.80NN Mobjack Bay to Rappahannock River

III 37.81-38.40NN Rappahannock River to Patuxent River

IV 38.41-38.80NN Patuxent River to South River/Annapolis

V 38.81-39.10NN South River/Annapolis to Bay

Bridge/Magothy River

VI 39.11-39.66NN Bay Bridge/Magothy River to Susquehanna

Flats
Nutrient Criteria—Estuarine and Coastal Waters G-25

Summary statistics on surface chlorophyll

concentrations (mg m-3) from Monitoring Program

cruises on the Chesapeake Bay, 1985-1990.
G-26 Nutrient Criteria—Estuarine and Coastal Waters

The Environmental Protection Agency Chesapeake Bay Monitoring stations in the

York River estuary.

York River
Nutrient Criteria—Estuarine and Coastal Waters G-27

Seasonal distributions of chlorophyll a and primary production in the York River system;

monthly means and standard errors were calculated from the 10 years data (1985-1994)

for chlorophyll a and from 7 years data (1988-1994) for primary production. Dashed line

at 10 μg 11-1 indicates our criterion for algal blooms and primary production shown in Fig

3F was measured at WE4.2.
G-28 Nutrient Criteria—Estuarine and Coastal Waters

Salinity dilution curves of DO, ammonium, orthophosphate, silicate, nitrite + nitrate,

chlorophyll a and light attenuation coefficient in the water column of the York River system

for low (1991), mean (1990) and high (1994)flow years during the summer-fall period (June

through October).
Nutrient Criteria—Estuarine and Coastal Waters G-29

Sampling locations (Stns 1, 5, and 6) in the lower Neuse River Estuary,

North Carolina, USA.

Neuse River Estuary, NC
G-30 Nutrient Criteria—Estuarine and Coastal Waters

Neuse River mean monthly flows (× 1000 ft3s

-1 or × 28.3

m

-3 s-1

), at the US Geological Survey gauging station in

Kinston North Carolina. Dashed line represents 60 yr

average flow (J.D. Bales, US Dept of Interior, Geological

Survey, Water Resources Division, Raleigh, North

Carolina). (Middle) Surface measurements of salinity at

Stns 1, 5, and 6 (see Fig. 1). Continuous line represents

means of stations measure. (Bottom) Monthly rainfall

totals at the Institute of Marine Sciences, Morehead City,

North Carolina. Dashed line represents average monthly

precipitation for the southern section of the AlbermarlePamilico estuarine system area (118.5 mm mo-1

) (H.

Porter, University of North Carolina Institute of Marine

Sciences, Morehead City).
Nutrient Criteria—Estuarine and Coastal Waters G-31

Nitrate, ammonium, and phosphate concentrations in

surface waters at Stns 1, 5, and 6 (see Fig. 1). 

Continuous line represents means among stations.
G-32 Nutrient Criteria—Estuarine and Coastal Waters

(Top and middle) Chlorophyll a concentration of selected

nutrient addition treatments minus controls, averaged over

the 4 d of each bioassay. N(14.3) and N(28.6) respectively

indicate addition of 14.3 and 28.6 μm NO3- P(3.2) indicates

addition of 3.2 μm PO4

3-. Error bars not visible are smaller

than symbol. (Bottom) Dissolved inorganic nitrogen:

dissolved inorganic phophorus (DIN:DIP) ratio (by atoms). 

No error bars plotted.
Nutrient Criteria—Estuarine and Coastal Waters G-33

(Top and middle) 14C assimilation of selected

nutrient addition treatments minus xontrols,

averaged over the 4 d of each bioassay. Symbols

as in Fig. 4 (Bottom Nitrate, ammonium, and

phosphate concentrations in surface waters at Stn 6

(see Fig. 1). Data are compiled from Fig.3 and

presented here for comparative puposes.
G-34 Nutrient Criteria—Estuarine and Coastal Waters

Sampling stations along the Cape Fear River Estuary, North Carolina,

United States. The lower estuary is centered in 33°56NN, 77°58NW.

Cape Fear River Estuary
Nutrient Criteria—Estuarine and Coastal Waters G-35

Physical parameters for selected stations in the Caper Fear River

Estuary, June 1995-November 1998.
G-36 Nutrient Criteria—Estuarine and Coastal Waters

Inorganic nutrient concentrations for selected stations in the Cape

Fear River Estuary, June1995- November 1998.
Nutrient Criteria—Estuarine and Coastal Waters G-37

Coastal Georgia

Nutrient exchange (:g-atom m-2 d-1) across the sediment/water interface and initial nutrient

concentrations (:g-atom 1-1) in the overlying water of the nearshore environment. Values

given as release (+) or uptake (-) of the nutrient by the sediment.

Constituent Initial Concentration :M

O2 6.5

NH4

+

1.1

NO2-0.11

NO3-0.16

DON 24.3

PO4/1.4

DOP 0.0

E Inorganic nitrogen 1.26

E N 25.56

E Inorganic phosphorus 1.4

E P 1.4
G-38 Nutrient Criteria—Estuarine and Coastal Waters

Map showing the distribution of estuaries in the Gulf of Mexico.

Gulf of Mexico Estuaries
Nutrient Criteria—Estuarine and Coastal Waters G-39

Site

Area

(km2)

Volume 

(106 m3)

Mean

Depth

(m)

Maximum

Length

(km)

Maximum

Width

(km)

Florida Bay, FL 1800 3200 2.0 70.0 40.0 Carbonate system low

freshwater input

Tampa Bay, FL 896 3490 3.9 61.1 16.1 Marine dominated estuary

Apalachicola Bay, FL 593 1600 2.7 12.9 33.8 River dominated with microtides

Mobile Bay, AL 1060 3200 3.0 48.5 40.0 River dominated with

seasonal coastal plume

Fourleague Bay, LA 56 72.8 1.3 16.0 3.5 Shallow, wetland dominated

salinity variable usually low.

Nueces River Estuary,

TX

538 1290 2.4 28.2 12.9 Event driven estuary, low

river flows

Celestun Lagoon,

Mexico

28 33 1.2 20.7 2.1

Freshwater spring-fed

lagoons.

Chelum Laggon,

Mexico

13.6 16.3 1.0 14.7 1.8

Dzilam Lagoon,

Mexico

9.4 11.2 0.8 12.9 1.6

Rio Lagartos Lagoon,

Mexico

96 76.8 0.8 80.0 1.5

Table 5-1. Physical Characteristics of Gulf Estuaries Examined in Case Studies.
G-40 Nutrient Criteria—Estuarine and Coastal Waters

Annual Mean and Range of Salinity and Nutrient Concentrations for Gulf Estuaries Examined in the Case Studies.

Annual mean and range of Chlorophyll-a Primary Production of Gulf Estuaries Examined in the Case Studies.
Nutrient Criteria—Estuarine and Coastal Waters G-41

Average Concentrations of Nutrients and Other Chemical Parameters in Galveston Bay, Calculated from the

TWC Database (± 1F).

Galveston Bay
G-42 Nutrient Criteria—Estuarine and Coastal Waters

Map showing location of Padilla Bay in relation to Puget Sound (left) and the study site (X) in

Padilla Bay.

Puget Sound

Padilla Bay
Nutrient Criteria—Estuarine and Coastal Waters G-43

Dissolved inorganic nutrient concentration in surface water at the

study site from June to October 1992.

DIN:PO4

3- ratios and chlorophyll a concentrations in surface water at

the study site from June to October 1992. DIN is NO3

-

 + NO2

-

 + NH4

+

. 

PO4

3- is soluble reactive phosphate. The dotted line represents the

Redfield ratio of 16:1.
Nutrient Criteria—Estuarine and Coastal Waters H-1

APPENDIX H

PRELIMINARY STATEMENT OF PROPOSED NEAR COASTAL MARINE

NUTRIENT SAMPLING AND REFERENCE CONDITION

DEVELOPMENT PROCEDURE

Synopsis of the National Nutrient Criteria Program, Coastal Marine Sampling Design Planning Meeting, 

4-5 June 2001, USEPA Environmental Science Center, Ft. Meade, MD.

Near Coastal Marine Nutrient Sampling and Reference Condition Development Committee Members:

Barry Burgan, USEPA; John Fox, USEPA; Laura Gabanski, USEPA; Jeroen Gerritsen, Tetra Tech Inc.;

George Gibson, USEPA*; William Muir, USEPA; Kent Price, Univ. Del.; Don Pryor, NOAA; Greg

Smith, GLEC, Inc.; Val Smith, Univ. Kansas; and Jack Word, MEC Analytical Systems, Inc.

*Contact: USEPA Laboratory, 701 Mapes Road, Ft. Meade, Md. 20755-5350. Phone: (410) 305-2618

E-mail: Gibson.George@EPA.gov.

Background and Purpose

Cultural eutrophication is an established water quality management concept and concern reaching as far

back as the 1600s in America (Capper, Power and Shivers 1983). However, extensive public recognition

of this form of pollution in coastal waters is relatively recent. The publication “Eutrophication, Causes,

Consequences, and Correctives” (NAS 1970) is often perceived as the technological beginning of

American nutrient pollution awareness and is centered on the understanding and abatement of this

problem primarily in freshwater lakes and reservoirs. We have since come to better understand the

problem in streams, rivers and estuaries with the publicity and public involvement in the Chesapeake Bay

studies of the 1980s. Vollenweider, Marchetti, and Viviani published “Marine Coastal Eutrophication”

in March of 1992 and this volume may be considered the coastal equivalent of the land mark NAS

freshwater publication a decade earlier. 

In response to this growing awareness, the EPA National Nutrient Criteria Program is preparing technical

guidance for nutrient reference condition determination and related criteria development to be used by

States and Tribes in the reduction of cultural eutrophication of the Nations’ surface waters. This report

concentrates on coastal marine waters and the effort to identify relatively natural nutrient water quality

conditions, which can be used as a benchmark to evaluate cultural eutrophication or overenrichment. A

preliminary literature investigation and data search indicate that insufficient data exist to derive the

reference condition information suitable for the needs of the Program without going to primary data

collection.

Because data gathering is likely to be a preliminary concern as well as an ongoing requirement, this

meeting of coastal marine nutrient research and management specialists was called to design a standard

protocol that the EPA can recommend for use in U.S. marine coastal waters. Coastal marine waters are

defined as those waters within 20 miles of shore along the East, West, and Gulf coasts of the United

States as well as Alaska, Hawaii, and the U.S. Trust Territories. Emphasis is on the three mile limit State

waters, although interest may devolve to the 12- mile U.S. limit as well. Nutrient loading from cultural

land run off sources are not presently expected to be a serious problem beyond this limit. The general

design and protocol are applied here to a case study example, the coastal waters of the mid-Atlantic Bight

from New Jersey to the Virginia Capes. Many of the design elements, in particular the number, size and

placement of spatial elements (strata and cells; see below) would need to be modified for specific

applications in different parts of the U.S. coastline.
H-2 Nutrient Criteria—Estuarine and Coastal Waters

A coastal transect of fixed stations exists for most of the Mid-Atlantic Bight, which has been used by

EPA and NOAA for several years to collect nitrogen, phosphorus, Secchi depth (SD), and chlorophyll-a

(Chl-a) data. This procedure and data base will be the prototype presented and discussed to develop the

recommended protocol.

Objective

To determine a simple, cost effective, scientifically defensible and standardized method to sample for

marine enrichment variables to use in determining reference condition for nutrient criteria derivation.

Premise of the National Coastal Nutrient Criteria Program

Offshore marine and onshore, near-coastal sites removed from point and estuarine discharges can be

identified as reference sites reflecting the least culturally impacted nutrient water quality of a region. 

“Region” in this case is a geographically similar portion of the coastline such as the Mid-Atlantic Bight. 

Such regions can also be described from the coastal portion of the Level III nutrient ecoregion map of the

continental United States (which is consistent with the rest of the National Program and is similar to the

ORD Provinces used by EMAP). 

Nutrient water quality is established from representative sampling of the coastal waters at these a priori

reference sites. The other elements of nutrient criteria, i.e., historical trends, modeling of the data for

additional insights, and attention to the consequences down-current of any proposed nutrient criteria, and

assessment of all of this information by a Regional Technical Assistance Group (RTAG) are applied to

the initial reference condition values to develop nutrient criteria for total phosphorus (TP), total nitrogen

(TN), Chl-a, and SD.

These criteria can then be used by States and Tribes to manage and monitor the nutrient quality of their

coastal marine waters. While this concept was developed and initiated by the USEPA beginning with the

Biological Criteria Program in 1989 (EPA-440/5-90-004, EPA-440/5-91-005) and further refined and

applied to nutrients by the EPA National Nutrient Criteria Program in 1995 (EPA 822-R-96-004, EPA

822-R-98-002), the idea has also been independently developed by the Swedish Environmental

Protection Agency using the reference condition approach on a regional basis and employing the same

indicator variables (Report number 5052, 2000).

Importance

All other waters of the continent drain to the coast and these coastal marine areas are the recipient of any

nutrients not intercepted from that cumulative runoff. Globally, conditions of many coastal areas have

shown several-fold increased levels of nitrogen and phosphorus since industrialization (Smith 1998,

Smith, Tilman and Nekola 1999), and preliminary assessments of empirical data collected from the MidAtlantic Bight between New Jersey and North Carolina by USEPA Region III since 1987 have suggested

an upward trend in the concentration of both dissolved inorganic nitrogen (DIN) and dissolved inorganic

phosphorus (DIP) at stations approximately 1 to 5 nm off shore (Muir, pers com, 2001). An extensive

baseline of region specific coastal nutrient data, regularly and consistently collected, is needed to

establish criteria against which future loading conditions may be compared. Potential predictive models

may also be developed from this information relating algal booms and other biological responses to these

nutrient levels in coastal waters.

The Transect Sampling Design

As illustrated below (Figure H-1, of coastal sampling stations between Atlantic City, NJ and Kitty Hawk,

NC) a series of transect sampling stations have been located between one and five miles from shore along

a designated coast line to measure ambient water quality as reflected in TN, TP, Chl-a and SD. Data

collected from those stations determined to be remote from significant cultural impacts such as sewage 
Nutrient Criteria—Estuarine and Coastal Waters H-3

Figure H-1. Example of an initial coastal monitoring project conducted off the mid-Atlantic

shelf just north of Delaware Bay and south of the mouth of the Chesapeake Bay to

obtain nutrient reference condition values. Open circles are reference sites and the

inshore references are compared to the offshore ones for confirmation.
H-4 Nutrient Criteria—Estuarine and Coastal Waters

discharges, industrial activities, major port facilities, or estuarine discharges constitute the reference

condition. The values measured can be compared to marine conditions further offshore (e.g., 20-25 nm)

reflecting, at least for TN and TP, the unimpacted condition. 

This comparison, together with attention to prevailing currents including upwellings, would establish the

seasonal values for TN, TP, Chl-a, and SD. Salinity would also be an important variable to document

local constancy of the waters and to avoid discharge plumes from rivers and estuaries that are not part of

the defined coastal waters of concern. These estuarine and riverine waters should have their own criteria.

Data have been collected from the surface 1 meter, mid depth, and the bottom 1 meter of the water

column at each station. Measurements are made on a seasonal basis, essentially mid-summer and midwinter and with sufficient data, criteria can be established for each season. While the protocol as

described above has been in use for about twenty years, most of those data are for dissolved inorganic

nitrogen and phosphorus rather than TN and TP, Chl-a, and SD. The TN, TP and off shore stations are a

recent addition and only two summers of data are available which is reported here.

Data Collection

Operations are during daylight hours only in order to include SD measurements with surface nutrient

collections and to maintain a consistent nutrient depth profile relative to photo periodicity. Additional

data collected with a CTD includes salinity, pH, temperature, depth, conductivity, and DO.

Data sampling points presently include the discharge plumes of estuaries, rivers, and point dischargers to

monitor impacts and design management plans, but not as part of the proposed coastal reference

condition sampling system per se.

The Pilot Project

The area approximately between Atlantic City, NJ and Kitty Hawk, NC has been studied by EPA Region

III for about 17 years and includes a mix of nutrient, chemical, biological, and physical measurements. 

This data has been processed and will provide trend information about the area.

For the last two summers and one winter, nutrient data have been collected from this area in the manner

described above. This is the initial basis for a reference condition determination and is presented in

Figures H-2 a-d, below. 

If the pilot project is judged successful, it is expected that the process, training, and funds for similar

equipment will be provided to each of the coastal EPA Regions for comparable operations to develop

their ecoregional coastal nutrient criteria. This area more than any other because of the proximity of

State and Federal waters will lend itself to joint data gathering and criteria development.

Methods and Materials

Forty-nine sampling stations are located along the 200 nautical mile (nm) stretch of coastal waters. 

There are twenty stations roughly 10 miles apart situated one to five miles off shore (total of 39

consisting of either single stations or sets of two or three), there are five intermediate (ten miles off

shore) stations, and there are eight stations located about 20 nm offshore (Figure H-1).

Sampling was conducted from the OSV Peter W. Anderson using a Sea-Bird brand CTD and rosette

sampler with 30 L Niskin bottles to produce a continuous water column profile and discrete water

samples from the surface one meter, the mid-depth, and the bottom one meter of the water column at each

station. One liter of sample was filtered using a Millipore Corporation apparatus and 0.7 um fiberglass

filters (Whatman GF/F). Ten ml sub-samples each of water were taken for TN and TP analysis using the
Nutrient Criteria—Estuarine and Coastal Waters H-5

0

0.01

0.02

0.03

0.04

0.05

 

TP (mg/L)

J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4

1999 5 NM offshore data 2000 5 NM offshore data 1999 & 2000 20 NM offshore data

Delaw are Bay Chesapeake Bay

 Reference Sites Plume Impacted Sites Referen. Sites Plume Impacted Sites Reference Sites

Figure H-2a. Two-year summer nutrient survey results using a sampling design as illustrated in

Figure 1. Potential reference condition for summer conditions is 0.025 mg/L TP

(NM = nautical miles).
H-6 Nutrient Criteria—Estuarine and Coastal Waters

0.1

0.15

0.2

0.25

0.3

0.35

 

TN (mg/L)

J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4

1999 5 NM offshore data 2000 5 NM offshore data 1999 & 2000 20 NM offshore data

Delaw are Bay Chesapeake Bay

 Reference Sites Plume Impacted Sites Referen. Sites Plume Impacted Sites Reference Sites

Figure H-2b. Two-year summer nutrient survey results using a sampling design as illustrated in

Figure 1. Potential reference condition for summer conditions is 0.175 mg/L TN

(NM = nautical miles).
Nutrient Criteria—Estuarine and Coastal Waters H-7

-1

0

1

2

3

4

5

6

 

Chlorophyll-a (ug/L)

J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4

1999 5 NM offshore data 2000 5 NM offshore data 1999 & 2000 20 NM offshore data

Delaw are Bay Chesapeake Bay

 Reference Sites Plume Impacted Sites Referen. Sites Plume Impacted Sites Reference Sites

Figure H-2c. Two-year summer nutrient survey results using a sampling design as illustrated in

Figure 1. Potential reference condition for summer conditions is 0.09 μg/L

chlorophyll-a (NM = nautical miles).
H-8 Nutrient Criteria—Estuarine and Coastal Waters

-2

0

2

4

6

8

10

12

14

16

 

Secchi depth (m)

J16 J17 J18 J19 CD11 J20 CD12 CD13 CD14 CD15 CD16 GG1 CD17 CD20 CC12 CC13 C14 GG2 GG3 GG4

1999 5 NM offshore data 2000 5 NM offshore data

Delaw are Bay Chesapeake Bay

 Reference Sites Plume Impacted Sites Referen. Sites Plume Impacted Sites Reference Sites

Figure H-2d. Two-year summer nutrient survey results using a sampling design as illustrated in

Figure 1. Secchi depth data are incomplete because of missing observations during

night-time operations (NM = nautical miles).
Nutrient Criteria—Estuarine and Coastal Waters H-9

Standard Methods persulfate digestion method. All samples were frozen on board for later analysis at the

University of Maryland Chesapeake Biological Laboratory. Secchi depth was determined using a 35 cm

diameter white Secchi disc on 0.5 m marked line. 

Results

Preliminary data analyses indicated that the groupings of two or three stations (originally intended to

reflect progressive off shore encroachment of nutrient runoff) are not significantly different. Similarly,

initial assessment of the surface, mid-depth, and bottom water column samples showed no significant

differences for the nutrient criteria indicators except for DO. Salinity did show the expected surface to

bottom variations expected of near coastal waters. Consequently horizontal and vertical sampling

distances for each station were combined to produce the nutrient results presented here. 

The following figures illustrate the results of two summer surveys in 1999 and 2000. Data indicated by

an asterisk represents the mean of combined two year data from each 20 nm off shore station. Gaps in

the Secchi depth results are because of the night time intervals during 24-hour operations. The stations

located in the vicinity of the estuarine discharges of Delaware and Chesapeake Bays are indicated by the

appropriately labeled open boxes, i.e. J18 through CD12 for the Delaware and CD17 through CC13 for

the Chesapeake.

Discussion

• The survey technique appears to faithfully reflect the nutrient conditions of the coastline in that

higher nutrients were found in the waters just off each estuary. Further, when surface station

salinity was plotted, a mirror image of the nutrient data results was presented indicating a

freshwater correlation with higher nutrient concentrations. The biological response of the waters

was also evident from the correlation of Chl-a and SD data with the TN and TP levels.

• Given the variability in only two years of data, it is interesting that the offshore stations appear to

be relatively consistent in all three nutrient parameters. The incongruity of the TP and TN data

at GG1 for 1999 and for chlorophyll-a at CD16 for 2000 is also noted, but unexplained.

• The two consecutive sampling years demonstrated similar trends among the stations indicating

no major weather variability during this period and also suggesting that the proposed reference

sites respond comparably to the discharge sites to inter-annual climatological events. 

• The stations that consist of sets of two or three sites within a few miles of one another were

determined upon review of the data to have very little distinction. This suggests that

perturbation, if any, originating from the coastal land mass has had an impact over all three sites. 

• It appears that the Delaware plume drifts more southward from the mouth of its bay than does the

Chesapeake. This may reflect the lower discharge volume of the Delaware and the influence of

the Hudson River discharge and Longshore current, both displacing the plume southward. 

Similarly, the slight northward offset of both the Delaware and Chesapeake plumes in 1999

relative to 2000 may be a response to changing coastal current dynamics from year to year.

• There was no significant difference between the eight offshore sites (20 nm offshore) and their

inshore counterparts. Further comparison of offshore stations to those counterparts within the

estuarine discharge plumes will help determine the sensitivity of the comparison. The

intermediate 10 nm stations do not appear to add substantial information and can be discontinued

except in the vicinity of discharge plumes to help define these margins.
H-10 Nutrient Criteria—Estuarine and Coastal Waters

While the data is limited, it is encouraging that the offshore controls are comparable to the

expected inshore reference sites. Candidate inshore reference sites are initially selected on the

basis of an apparent physical absence of local cultural impact, i.e. tributary discharges, municipal

discharges, ports or marinas, or other commercial enterprises. Because of the potentially high

variability in the existing data, a concurrence between offshore and inshore reference stations is

not necessarily confirmation of the quality of the inshore reference sites, but when consistent 

with observed physical indications, this information adds confidence to the selection. 

Conversely, significant differences would be cause for suspecting the inshore site selection as of

reference quality. 

• In this regard, an interesting trend was noted among the three groups of reference sites located

north of Delaware Bay; between Delaware and Chesapeake Bay; and south of Chesapeake Bay

(Figures 2a-d). The mean ambient concentrations of TP, TN and Chlorophyll-a at these

reference sites trend downward from north to south. The same trend appears in the eight

stations 20 nm offshore presumably reflecting a broad scale process affecting this area. This

further supports the importance of establishing relatively close spaced reference sites when

preparing coastal marine criteria The mean TP values for the nearshore reference sites at the

northern terminus of the transect of stations is significantly higher than those at the southern end

(p = 0.0006) even though the region is presumed to be geologically homogeneous. 

• Secchi depth data were inconclusive because not enough data points were generated as a

consequence of 24-hour sampling when Secchi depth could not be determined during night-time

hours. Additional future sampling will be conducted during daylight so all parameters may be

evaluated. 

Committee Discussion of the Prototype Methodology

Fixed Station Sampling vs Stratified Random Sampling

Inferences derived from fixed-station and fixed-transect sampling, while common in oceanographic

research and monitoring, are potentially confounded by unintentional and unknown biases, and by the

inability to extend statistical inferences to the entire sample space desired. Alternatively, fixed stations

tend to reduce the amount of unknown physical variability associated with interpreting climatic factors

upon a given site such as when attempting to assess hurricanes, upwelling or acid deposition effects on a

particular coastline. Although there is no reason to suspect that the existing stations of the mid-Atlantic

coastal nutrient study are biased, the design group thought that the design should allow data inference to

the entire coastal sampling space. It therefore proposes a change to a probability-based design, of equal

sampling cost, to avoid the potential pitfalls of a fixed-station design.

The sample space for this project is open marine waters of the U.S. coastal zone, with emphasis on state

waters within the 3 nautical mile state limit. Sampling will be carried out in three sampling strata for

which nutrient conditions are to be estimated:

1. Reference areas within the 3-mile limit, outside the influence of major estuary plumes (e.g.,

Hudson River, Delaware Bay, Chesapeake Bay);

2. Nutrient influenced areas within the 3-mile limit, affected by the estuary plumes and other

discharges; and

3. Offshore waters beyond the state 3-mile limit.
Nutrient Criteria—Estuarine and Coastal Waters H-11

These three regions will define the sampling strata. The sampling design will be to define longshore

“cells” in each of these three zones. During each sampling event, one site will be selected randomly

within each sampling cell. The overall design can be described as stratified-systematic-random; where

the strata are the three areas defined by estuary influence and distance from shore, the systematic

component is the cells that define each stratum, and the random component is the random sampling

location selected on each cell during each sampling event (Figure H-3).

The first task in developing the sampling plan will be to define areas of presumed estuarine influence,

from existing physical oceanographic research on water movement, in this case in the mid-Atlantic

coastal area, and from existing water chemistry data showing elevated nutrients and other constituents

such as salinity and conductivity in the estuary plumes. Because plumes will vary with estuarine

discharge, Gulf Stream eddies, and other events, precise definition of the plume area is not possible. 

Instead, plume areas should be defined as where the estuarine influence is likely to occur. An

understanding of local estuarine hydrology will help in this regard, but extensive and expensive physical

investigations should not be a prerequisite for the determination of likely plume influence.

Coastal reference site determinations and reference condition derivation should be, at least initially,

established by EPA. Data collection would be continued as an EPA function because the offshore

stations are the purview of the Federal government and because the related reference sites can be

incorporated in the effort. 

Estuarine and riverine plume monitoring are more likely accomplished by the coastal States using

existing budgets and vessels already at their disposal. A coordinated effort relating State and Federal

sampling and data exchange should be promoted. The nutrient quality of the discharge plumes can then

be expeditiously compared to the proximal reference condition(s) to assess impact upon the near coastal

marine waters. 

Regional coastal characteristics will determine both the reference site cell structures and placement and

the estuarine or riverine plume sampling designs.

Sampling Times and Depths

The committee recommended that sampling be conducted during the period of optimal marine vegetative

growth. In temperate areas this is Spring and Summer, generally May, June and July, August. Other

interval options depending on locale might be wet and dry periods during the growing season.

As a cost effective approach producing scientifically valid results, sampling depth is recommended as

always at the surface, i.e., top meter of water accompanied by either a composite sample from the

remainder of the water column or sampling from just below the thermocline and at one meter above the

bottom. Some members of the committee advocate the surface, mid-depth, and bottom sampling

technique where the mid-depth sample is usually below the thermocline and/or where inshore waters

often fail to demonstrate a thermocline.

Variables

The primary four variables of TN, TP, Chlorophyll-a, and SD are recommended because they are the best

early indicators of causal and biological response indicators to nutrient loadings. Other measures of

clarity or transparency may also be used, but Secchi depth should always be included because such a

large body of information is already available in this form; it is inexpensive and reliable; and continued

Secchi depth measurements provide a continuity with much historical data. In an independent
H-12 Nutrient Criteria—Estuarine and Coastal Waters

U U

U U

ll l

l

l

l

l

l

l

l

10 N M 20 N M

10 N M 20 N M

1-3 N M

A

B

C

D

E

Figure H-3. Illustration of stratified random grid sample design for reference condition cells as

related to an estuarine discharge. Nutrient quality and spatial extent of the grid

can be compared to a value derived from the measurements in reference sites (cells)

A-E. Values at 10 and 20 nm stations are a check against reference values found at

A-E. 
Nutrient Criteria—Estuarine and Coastal Waters H-13

investigation, the Swedish Environmental Protection Agency selected the same four primary variables

(Report 5052, 2000).

Other recommended variables are: dissolved oxygen as an important secondary response variable almost

always measured by investigators because of the significance of respiration to the biological community,

and planktonic species composition as a refined diagnostic indicator of the nature and extent of

enrichment. 

Geographic Application of the Protocol

Members of the committee are familiar with both the East, West, and Gulf coasts of the continental U.S.

and conclude that the method described above, with allowances for regional modifications such as

relative distance from shore to shelf break and the magnitude of upwelling, can be successfully applied in

all three coastal environments to identify reference conditions for criteria development.

Summary Conclusions

1. The basic protocol as described, but modified to include a probability-based sampling design, a

variation on the surface, mid-depth, and bottom sampling profile, and sampling emphasis on

twice during the growing season, is a scientifically defensible and broadly applicable method for

establishing regional reference conditions to support coastal marine nutrient criteria

development. 

2. The TP, TN, Chlorophyll-a, and perhaps Secchi depth variables are responsive and together with

salinity measurements are descriptive of both estuarine discharge plumes and near coastal

reference quality waters. These measurements can be used to assess the concentration and area

changes of discharge plumes over time.

3. The comparison of marine nutrient water quality to inshore reference sites is valuable as

confirmation of “natural” reference conditions and as a graphic descriptor of the reference

concept for the public. However, for cost effectiveness, these off shore stations need not be

monitored every time the inshore reference sites are sampled.

4. But it is important to note that the distinction of cultural from inherent nutrient discharges by

estuaries and other tributaries to coastal waters is not possible just by comparison to reference

conditions. The other elements of nutrient criteria development should be incorporated to help

make this distinction. The reference conditions and criteria will reveal exceedences of the

“natural” background levels to be preferred and the relative extent and magnitude of the problem,

but source identification and cause and effect studies will be required for an effective

management response to this identified concern. 
Nutrient Criteria—Estuarine and Coastal Waters R-1

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